1 OBJECTION TO PEEL ENERGY LIMITED PLANNING

Transcription

OBJECTION TO PEEL ENERGY LIMITED PLANNING APPLICATION FOR A
BIOMASS RENEWABLE ENERGY PLANT AT INCE
CWaC PLANNING REFERENCE 11/00040/WAS
I, Professor John C. Dearden, of 10 Landscape Dene, Helsby, Cheshire WA6 9LG, hereby
object to the above planning application, for the reasons given below.
I am Emeritus Professor of Medicinal Chemistry at Liverpool John Moores University, where
I have worked for the last 46 years. I hold a B.Sc. in chemical engineering, a M.Sc. and a
Ph.D. in physical organic chemistry, and an Associateship of the City & Guilds Institute; I am
also an honorary member of the Royal Pharmaceutical Society of Great Britain, for
contributions to pharmaceutical research. My prime area of work is in computational
toxicology, and I was the 2004 r ecipient of the biennial International QSAR Award for
Research in Environmental Toxicology. I am the author of over 250 scientific publications in
computational toxicology and related fields. I serve on a European Commission working
party in connection with the recent REACH (Registration, Evaluation and Authorisation of
Chemicals) legislation, and in 2001 I was invited to give evidence to the Royal Commission
on Environmental Pollution.
1. Lack of need
Policy EM17 of the North West of England Plan: Regional Spatial Strategy states that there is
a requirement in Cheshire for 4MW of biomass generating capacity by 2010, rising to 9MW
by 2015 and 2020. However, Cheshire’s current biomass energy generating capacity is about
8MW, according to Peel’s Planning Statement para. 8.9. Furthermore, Fiddlers Ferry power
station already has the capacity to generate 200MW from biomass. Hence there is no
justification for a plant that would produce 20MW of electricity and with the potential
to produce 5MW of heat.
Peel’s Planning Statement para. 8.11 s tates that the generation of electricity from wind and
wave in the North West is currently 341.1MW, 56.9% of the 2010 target; the same paragraph
goes on to state that: “Accordingly, these development proposals will go some way towards
making up for the current shortfall in achieving the targets of the NWRSS across other
renewable technologies”. I submit that this argument is nihil ad rem, an irrelevance. Biomass
electricity generating capacity cannot contribute towards generating capacity requirements
for wind and wave.
2. Lack of carbon neutrality
Peel claim that the incinerator would be carbon-neutral, because it would simply release back
to atmosphere the carbon dioxide (CO 2 ) that the biomass absorbed whilst it w as growing.
This is a fallacious argument, for five reasons:
(i)
(ii)
since wood grows slowly, it would take many years before the CO 2 emitted by the
incinerator was re-absorbed (a generally accepted figure for the half-life of
atmospheric CO 2 is about 40 years, and a recent study [1] has found that it takes
about 40 years for a biomass energy plant to become carbon-neutral, whereas the
life-time of Peel’s proposed plant is stated to be 25 years);
if the wood was recycled instead of being burned, the CO 2 emissions would be
much lower, because wood-processing (e.g. to make chipboard and other wastewood products) releases about 380 kg of CO 2 per tonne of wood processed, in
contrast to incineration of wood which releases about 1900 kg of CO 2 per tonne of
wood processed. The difference (1520 kg of CO 2 per tonne of wood processed)
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means that the proposed Peel biomass incinerator would emit up to an extra
268,000 tonnes of CO 2 per annum, compared with recycling;
Peel ignore the CO 2 emissions of transportation of biomass to Ince. Peel’s nontechnical summary of their Environmental Statement states that there would be a
maximum of 36 two-way HGV journeys per day. Peel also state that
“approximately 70% of the biomass fuel will consist of waste wood that will
primarily be sourced from local markets in the north west region”. The other 30%
would be virgin wood. Assuming that an average transportation distance would be
50 miles (100 miles there and back), the HGV mileage per annum would be
1,314,000. Average HGV fuel consumption is 8 mpg, so annual diesel used would
be 164,250 gallons, which would release about 2000 tonnes of CO 2 .
According to Peel’s figures, about 53,000 tonnes per annum of virgin wood would
be burned in the biomass incinerator. Virgin wood is wood that has not been
treated, cut or shaped in any way, and can include logs, bark, arboricultural
arisings, sawdust and wood chips. However, removal of such waste material from
forests, instead of allowing it to rot naturally, weakens the carbon sink capacity of
forests, and thus reduces the capacity of forests to absorb atmospheric CO 2 ,
according to a report from the Finnish Environment Institute [2]. It thus in effect
increases atmospheric CO 2 levels. In addition, the CO 2 and other emissions
generated by the removal of virgin wood have not been taken into account [2].
Biomass energy plants are among the worst polluters in terms of CO 2 . Based on
the figure of 1900kg of CO 2 per tonne of wood burned, the proposed plant would
emit 335,350,000 ( 1900 x 176,500) kg pe r annum. The electrical output of the
plant would be 20MW, and according to Peel’s Non-Technical Summary the plant
would operate for 90% of the time. Hence the annual output would be 157,680 (20
x 8760 x 0.9) MWh. The CO 2 emissions would therefore be 2127 kg/MWh.
(iii)
(iv)
(v)
U.S. Department of Energy figures for CO 2 emissions from other types of
electricity generating plants are [3]:
gas 596kg/MWh; oil 867kg/MWh; coal 960kg/MWh.
2500
2000
1500
1000
500
0
Gas
Oil
Coal
Biomass
Figure 1. CO 2 emissions (kg/MWh) from power plants
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Clearly the biomass plant would be by far the worst polluter in terms of CO 2
emissions. Even if one takes into account the 5MW of heat that might be used, the
CO 2 emissions from the proposed Ince biomass plant would still be by far the
highest at 1702kg/MWh. I submit that this level of CO 2 pollution is
unacceptable.
In January 2011 Peel Energy’s website (www.peelenergy.co.uk) compared the parameters of
the bio-ethanol and biomass plants:
Item
Process
By-products
Stack height
Bio-ethanol plant
Chemical
CO 2 , lignin fuel, distillers grain
30 metres
Biomass plant
Combustion
Ash, steam, hot water
85 metres
Note that Peel did not say that the biomass plant would emit CO 2 . Their non-technical
summary similarly contains the same omission, as did their public presentations to local
authorities. I consider this to be a deliberate attempt to play down the fact that the plant
would emit about 335,350 tonnes of CO 2 per annum.
3. The waste hierarchy
Annexe C of the Government’s Planning Policy Statement 10 ( PPS10) presents the waste
hierarchy, which defines the Government’s preferred options for waste treatment as: 1.
Reduction; 2. Re-use; 3. Recycling & Composting; 4. Energy Recovery; 5. Disposal (usually
landfill). Waste management should always aim to be as high as possible in the waste
hierarchy.
Peel’s biomass incinerator proposal falls in the next to lowest level of the waste hierarchy.
Since wood can viably be recycled, the proposal goes against the waste hierarchy, and thus
conflicts with one of the core objectives in PPS10. PPS10 § 23 makes it clear that Waste
Planning Authorities (WPAs) should determine proposals in a way that is consistent with the
policies of PPS10, and PPS10 § 3 states that WPAs should deliver strategies that “help
deliver the national waste strategy and supporting targets”.
The Government’s Waste Strategy for England (2007) promotes adherence to the waste
hierarchy. Its Annexe K states that: “In particular, WS2007 makes clear that energy should be
recovered only from residual waste that cannot viably be recycled”. Peel’s proposal thus
conflicts with WS2007 as well as with PPS10. By going against the waste hierarchy, a grant
of permission for Peel’s biomass planning application would conflict with the policy
objective enshrined in PPS10, and therefore this application should be refused.
Page 9 of WS2007 includes a diagram of the waste hierarchy and goes on to note that: “The
dividends of applying the waste hierarchy will not just be environmental. We can save money
by making products with fewer natural resources, and we can reduce the costs of waste
treatment and disposal. Waste is a drag on the economy and business productivity. Improving
the productivity with which we use natural resources can generate new opportunities and
jobs.”
Paragraphs 20 and 21 of chapter 1 of WS2007 make it clear that adherence to the waste
hierarchy encapsulates the Government’s overall objectives for waste policy. The waste
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hierarchy outlined in Figure 1.3 of WS2007 is consistent with the waste hierarchy in PPS10,
and makes it clear that recycling is to be preferred over energy recovery.
Chapter 5 of WS2007, § 1 states that: “We need waste to be minimised to the greatest extent
practicable, and such waste as does arise to be managed as far up the waste hierarchy as is
reasonably achievable. Resources should be recovered in ways that maximise the costeffective reduction in greenhouse gas emissions over the lifecycle.”
Chapter 1 of WS2007, § 22 s tates that: “Recent studies have confirmed that the waste
hierarchy remains a good general guide to the relative environmental benefits of different
waste management options...”
As recycling is higher on the waste hierarchy than energy recovery then Peel should be
required robustly to demonstrate that either the timber would otherwise go to landfill or that,
in this specific instance, the waste hierarchy is mistaken.
Peel’s application fails to demonstrate that either:
(a) energy recovery would be only from residual waste that cannot viably be recycled; or that
(b) there are clear carbon benefits in using the timber for biomass energy recovery, instead of
as material available for wood panelling (see below).
PPS1 Supplement on Climate Change makes clear that “tackling climate change is a k ey
Government priority for the planning system”. Thus Peel’s proposed biomass plant conflicts
with WS2007 and PPS1 Supplement on Climate Change, and should therefore be refused
planning permission.
It should be noted that early in 2011 the Planning Inspectorate rejected an appeal by Resource
Recovery Solutions for a 190,000 t onne/year gasification and mechanical biological
treatment (MBT) plant in Derby on t he grounds that it ignored the waste hierarchy. The
Planning Inspector said: “This facility’s appetite for waste could divert efforts and resources
away from the promotion and encouragement of waste reduction, re-use and
recycling/composting”.
Wood Panelling
Wood panels are produced using a range of sources, including small roundwood, chips,
sawdust and recycled wood. Each type of panel has various applications in the construction,
furniture and do-it-yourself sectors, including cladding, packaging, kitchen worktops and
laminate flooring. Panels are vital components that can be replaced only by more expensive
and less sustainable products.
Subsidies that encourage the use of wood harvest in energy generation are causing tension
between the processing industry and the energy sector. The direct use of biomass and wood
for energy production is not only reducing the wood supply but is also creating negative
consequences for the environment.
The hierarchy of use principle would help rationalise the use of wood and define preferred
options, i.e. using and recycling wood, with burning only as a last resort. If more and more
wood and forest residues go directly to energy plants we are wastefully minimising the
carbon cycle of wood.
A 2010 report entitled Carbon emissions for end of life scenarios for wood fibres [4] was
commissioned by the Wood Panel Industries Federation. This report, which is primarily
concerned with net CO 2 emissions arising from competing uses for the U.K.’s scarce and
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finite timber resources, maps and compares the emissions associated with the processing of
one tonne of wood through the wood panel production process and the transport and eventual
burning of one tonne of wood to generate electricity.
The report explains how, until recently, the wood panel industry has sourced its timber
domestically in the UK on a competitive basis. The competitive market began to be
undermined in 2002 by the Government’s introduction of a subsidy to compensate electricity
generators for burning “renewable” fuel in the form of the Renewables Obligation Certificate
(ROC). It is important to note also that this is also forcing up prices in manufacturing and
construction industries [5]. A spokesman for furniture manufacturer Senator International, the
largest manufacturer of office furniture in the U.K., recently stated: “Biomass burning wood
is hitting our industry and any industry that uses wood-based products”. This point is
reinforced in a r ecent report by Europe Economics [6]. Also, a P arliamentary debate on 16
March 2011 [7] highlighted the problems that wood biomass burning is causing for the wood
products industry, in particular because the large subsidies available for wood biomass
burning are distorting the market; the Energy Minister, Greg Barker, responded by saying: “I
am very aware of unintended or perverse consequences. We will work harder to look at the
consequences for the wood panel industry. Many powerful arguments were made today, not
least how it is better to lock up carbon rather than to burn it (my emphasis), and I am
mindful of that”.
At § 1.5 of Carbon emissions for end of life scenarios for wood fibres it is stated: “As a result
of the ROC subsidies, the projected demand for timber in the UK will outstrip supply by
2012. The biological availability of British sourced wood fibre is forecast to increase up t o
about 2019 when it reaches just over 20 million tonnes per annum and then it is forecast to
start decreasing. Demand during the same period is set to increase to 50 million tonnes as a
result of proposed increases in biomass electricity generation.” In Scotland alone there are
plans for four huge biomass energy plants [8] that between them would consume at least 5.3
million tonnes of wood per annum – the equivalent of almost two thirds of the U.K.’s annual
wood production of 8.4 million tonnes. In Cheshire, Fiddlers Ferry already has the capacity to
generate 200MW of electrical power from biomass, which equates to almost 2 million tonnes
of wood per annum.
4. Toxic emissions
(a) Toxic chemicals. Even virgin wood can produce toxic emissions from combustion, such as
dioxins and polycyclic aromatic hydrocarbons (PAHs), both of which are carcinogenic. Some
materials used to treat wood, such as chromated copper arsenate and creosote, are also very
toxic. Old painted wood may contain lead. The EU Hazardous Waste Directive states that
many wood treatments may contain materials that can cause undesirable or dangerous
emissions. Peel’s documentation lists, in addition to the above, the following pollutants that
would be emitted to atmosphere: nitrogen dioxide (NO 2 ), sulphur dioxide (SO 2 ), carbon
monoxide, hydrogen chloride (HCl), hydrogen fluoride (HF), and volatile organic compounds
(VOCs). It should be noted that NO 2 , SO 2 , HCl and HF can cause respiratory problems, and
NO 2 is a greenhouse gas 300 times as potent as CO 2 .
No indication is given in Peel’s documentation as to how, if at all, attempts would be made to
avoid burning contaminated wood. Clearly it would be impracticable to examine, test and sort
every piece of wood, and thus the likelihood is that contaminated wood would be burned in
the Ince biomass plant. Hence there would be, in my view, a significant risk of airborne
incineration products exceeding permitted emissions levels.
In Peel’s Environmental Statement, predicted emissions of a range of pollutants are given for
the biomass plant alone (Table 5.14) and for the whole of the Resource Recovery Park
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including the biomass plant (Table 5.20). It can be seen that the biomass plant would be
responsible for up to 30% of the levels of pollutants from the whole Park, which cannot be
dismissed as insignificant against the AQO/EAL limits for nitrogen dioxide (NO 2 ), volatile
organic compounds (VOCs) and polycyclic aromatic hydrocarbons (PaHs). S uch a large
increase, and the associated health concerns, are unacceptable when set against the dubious
requirement for a biomass plant in the broader context of carbon neutrality and waste
hierarchy.
In Peel’s Environmental Statement there appear to be mathematical errors which have
resulted in underestimation of the long-term levels of some heavy metals. T hese call into
question the conclusions about impact on air quality of the biomass plant, either alone or with
the Ince Park development. In Table 5.17, the concentrations of antimony, arsenic,
chromium, cobalt, copper, lead, manganese, nickel and vanadium have, according to section
5.5.26, been calculated assuming that each metal contributes one ninth of the combined limit.
The annual mean value for these “other metals” provided by the model and shown in Table
5.14 is 1.82 ng /m3. T hus the contribution of each individual metal is 0.2022 ng/m3 (1.82
ng/m3 ÷ 9), and not 0.0202 ng/m3 as shown in Table 5.17. If this is indeed an error, the
corrected level of arsenic is 6.7% of EAL and thus cannot be screened out as insignificant.
Also, despite the information in sections 5.5.31 and 5.5.32, the calculated long-term value for
chromium (VI) presented in Table 5.17 i s not clear, and I am concerned that there may be
mathematical errors leading to underestimation of that metal.
The same miscalculation in annual values for heavy metals appears to have been made in
Table 5.22, and the same comments apply.
I wish to stress in particular the dangers of dioxin emissions. Dioxins are among the most
toxic chemicals known, and according to the U.S. Institute of Medicine can cause cancers,
diabetes, nerve disease and heart disease in people exposed directly or indirectly, and can
cause spina bifida in their children.
Dioxins are a family of 75 polychlorinated dibenzo-p-dioxins (PCDDs), the structure of the
most toxic of which (TCDD) is shown in Figure 2. This compound is one of the most toxic
chemicals known, and is a known human carcinogen and endocrine disruptor. Similar
chemicals are polychlorinated dibenzofurans (PCDFs), of which there are 135 (see Figure 2).
Other related compounds are polychlorinated biphenyls (PCBs), of which there are 209,
many of which are known [9] to be endocrine disrupters; the structure of unsubstituted
biphenyl is shown in Figure 2. Yet others are polybrominated diphenyl ethers (PBDEs), of
which there are 209, and polybrominated biphenyls (PBBs) of which there are also 209; the
structure of unsubstituted diphenyl ether is shown in Figure 2. PBDEs and PBBs are used as
flame-retardants for electrical goods, clothing and furniture. They are known to be endocrine
disruptors and to cause developmental neurobehavioural defects [10, 11]. The principal cause
of their presence in the environment is widely accepted to be incineration [12].
All these compounds are hydrophobic (lipophilic) and therefore tend to accumulate in
adipose tissue in the body. They are also chemically very stable and are therefore resistant to
metabolic attack, and therefore to excretion, since chemicals need to be reasonably soluble in
water in order to be readily excreted. For example, the half-life of dioxins in humans is about
seven years.
PCBs, PBBs and PBDEs can be present in waste materials. Dioxins (PCDDs and PCDFs) are
not normally present in waste, but are formed when chlorine-containing organic substances
(e.g. PVC and wood) are burned. If combustion takes place at temperatures of about 850ºC,
any dioxins already formed are destroyed, but can re-form again post-combustion. Cunliffe
6
9
1
O
8
2
7
3
O
6
4
5
Dibenzo-p-dioxin, showing the 8 positions that chlorine can occupy (1 – 9, excluding 5)
1
Cl
9
O
Cl
8
2
3
7
Cl
O
5
4
Cl
6
2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD), the most toxic chlorinated dioxin
1
9
2
8
7
3
O
6
4
5
Dibenzofuran, showing the 8 positions that chlorine can occupy (1 – 9, excluding 5)
2
1
6
7
3
8
4
5
10
9
Biphenyl, showing the 10 positions that chlorine can occupy
2
6
1
7
O
3
4
5
8
10
9
Diphenyl ether, showing the 10 positions that bromine can occupy
Figure 2. Molecular structures of dioxins, furans, PCBs and PBDEs
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and Williams [13] found that “formation of PCDD/PCDF on f ly-ash deposits in the postcombustion plant of incinerators can result in the release of significant amounts of
PCDD/PCDF to the flue gas stream”. Littarru [14] has shown that about 57% of emitted
dioxins (in terms of TCDD equivalents) are in the flue gases, with about 43% sorbed on the
fly-ash.
In 1997 Douben [15] of H.M. Inspectorate of Pollution stated that “MSW incinerators are the
dominant source of PCDD/F emissions to atmosphere and are responsible for up to 80% of
the inventory”. It is now acknowledged that dioxin emissions from incinerators have fallen
considerably in recent years. However, there remain a number of areas of concern.
1. Dioxin emission levels from incinerators are measured once or twice a year by
external assessors who have to give prior notice of their visits. It is thus likely that
operators ensure that a plant is running under optimal conditions for a visit. If much
more frequent or continuous measurements are made, the total dioxin emissions are
found to be very much higher than those calculated from biannual measurements. De
Fré and Wevers [16] found that emissions measured using the European standard
method EN 1948 over a 6-hour period were 30 to 50 t imes lower than the average
over a t wo-week continuous period. Reinmann et al [17] showed that use of
continuous dioxin sampling enabled operators to reduce dioxin emissions by a factor
of 10, through careful control of operating conditions. True dioxin emissions from the
proposed Ince biomass plant, which would be subjected only to biannual checks,
would probably be very much higher than claimed.
2. Incinerators do not , for various reasons, run under optimal conditions all the time.
Grosso et al [18] found that even under steady-state conditions total dioxin release
varied between 1.5 and 45 µg TEQ per tonne of waste burned, depending on whether
activated carbon was used and how fly-ash was collected. S am-Cwan et al [19]
investigated the post-combustion re-synthesis of dioxins, and found that levels at
waste heat boiler outlets were 10.8 – 13.6 times higher than at the furnace outlets.
Incinerators have to be shut down on oc casion, both for routine maintenance and
because of operating problems. It has been observed that during shutdown and startup,
the levels of dioxins and other pollutants can be much higher than under optimal
operation. Tejima et al [20] tested the dioxin stack emissions of an MSW incinerator
under conditions of startup, steady state and shutdown. They found concentrations of
WHO-TEQ dioxin of 36 – 709 µg.m-3 during startup, 2.3 µg.m-3 during steady state
operation, and 2.5 – 49 µg.m-3 during shutdown. They estimated that 41% of the total
annual emissions could be attributed to the startup period, assuming three startups per
year. L.-C. Wang et al [21] found that a single start-up could contribute about 60% of
the PCDD/F emissions for one whole year of normal operations; hence, assuming
three startups per year, 64% of total annual emissions could come from startup. H.C.
Wang et al [22] found that during startup the PCDD/F removal efficiency was only
42% with selective catalytic reduction, compared with > 99% during normal
operation.
It is clear from the above that levels of pollutants emitted from incinerators can vary
greatly, and can exceed the statutory limits placed upon t heir emission. (It must be
noted here that those limits are generally based on what is achievable and measurable,
rather than what is safe, as was pointed out in the House of Commons Environment,
Transport and Regional Affairs Committee 5th Report in March 2001). In 2003 there
were four incidents of dioxins and furans above permitted levels, and one incident of
cadmium emissions above permitted levels. In 2010 two waste incinerators, one on
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the Isle of Wight and the other at Neath Port Talbot, were shut down by the
Environment Agency for persistent exceedances of permitted dioxin levels.
3. Incineration produces two forms of solid residue – fly-ash, which is fine particulate
matter carried with flue gases, and bottom ash, which falls from the fire-grate. They
constitute, between them, about one quarter to one third of the total pre-combustion
weight of waste.
Fly-ash is known to sorb chemicals from the flue gases. As pointed out earlier, around
half of emitted dioxins are sorbed on fly-ash [14]. Fly-ash is also responsible for the
so-called dioxin memory effect [13], whereby slow de novo synthesis of dioxins
occurs on the surface of the fly-ash; the dioxins then slowly desorb into the flue gases
[23] for prolonged periods after the implementation of beneficial changes to the
incineration process. Fly-ash is classed as hazardous waste, and has to be disposed of
to landfill. There is concern that, because of its dust-like nature, less than extremely
stringent handling could disperse dioxins and other pollutants such as metals sorbed
on the fly-ash into the atmosphere around the Resource Recovery Park.
Bottom ash contains similar proportions of heavy metals (except cadmium, which is
lower than in fly-ash). Under the List of Wastes (England) Regulations 2005,
incinerator bottom ash is classed as non-hazardous. However, the Environment
Agency recently confirmed, in a letter to Mr. Alan Watson of Public Interest
Consultants [24], that 12 out of 96 bottom ash samples that they tested met the criteria
for hazardous waste. This probably means that all of the bottom ash from the biomass
plant would have to be disposed of as hazardous waste, and should not be used for
block-making or indeed for any other purpose, which would mean a significant
financial disincentive for biomass incineration.
All of the above suggest that the dioxin emissions from the proposed Ince biomass plant
would be many times those claimed in Peel’s Environmental Statement.
A typical daily intake of dioxins is 1-4 picograms per kilogram of body weight per day (14pg/kg.day). That is currently considered acceptable, although the U.S. Environmental
Protection Agency (EPA) is about to revise those figures downwards very markedly. A
particular problem arises with unborn foetuses and newborn babies, whose systems do not
have the ability to protect them against injurious chemicals, and who thus can be damaged
irreversibly. Foetuses absorb dioxins from their mothers via the placenta, and newborn babies
via their mothers’ milk. The Parliamentary Committee on Toxicity of Chemicals in Food,
Consumer Products and the Environment 2001 Report [25] stated that: “In infants under two
months of age, the estimated intakes of dioxins and dioxin-like PCBs from breast milk could
result in daily intakes of 20-60 times the TDI”. By the time infants’ protective mechanisms
have developed, the damage is done. This is why the argument is fallacious that the high
doses received in utero and in the first few months of life average out safely over a lifetime.
Recent Dutch research has shown that foetuses and neonates whose mothers are exposed
even to so-called background levels of dioxins develop a range of problems, including birth
defects, decreased lung function, persistent haematological and immunological disturbances,
delayed puberty, and dental malformations [26, 27, 28, 29, 30 ]. Thus there appears to be no
discernible threshold below which dioxins pose no health risk. But worse than that, recent
Finnish research [31] has shown that the carcinogenic effect of dioxins actually increases at
low doses, a phenomenon termed hormesis.
No information has been provided about the effects of dioxin emissions from the proposed
biomass plant on “sensitive receptors”.
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For the sake of the children of this region in particular, we must not allow any more
dioxin-producing industrial plants to be built in north Cheshire.
(b) Particulate emissions. Like all incineration, biomass burning produces particulate
emissions. Some of these are filtered out, but some (especially the smallest particles, which
are the most dangerous) cannot be completely filtered out. These particles can penetrate deep
into the lung and cause serious problems. In June 2009 the American Lung Association wrote
[32] to the U.S. Government to say that the Association “urges that legislation not promote
the combustion of biomass. Burning biomass could lead to significant increases in emissions
of nitrogen oxides, particulate matter and sulfur dioxide and have severe impacts on t he
health of children, older adults, and people with lung diseases”. A 2010 House of Commons
Environmental Audit Committee report on air quality [33] stated that: “In a number of cases
the climate change agenda has resulted in measures that increase air pollution. An…example
has been the promotion of biomass boilers in urban areas already suffering poor air quality”.
It is now accepted that particulate air pollution is a very significant cause of illness and
mortality [34, 35, 36]. Airborne particles are classified according to their size. Particles with a
diameter of ≤ 10 microns (1 micron (µm) = 1 0-6 metre) are potentially dangerous because
they are small enough to be drawn into the lung; such particles are designated PM 10 s.
Particles with a diameter of ≤ 2.5 µm are more dangerous because they can be drawn deeper
into the lung; they are designated PM 2.5 s. Even smaller particles are even more dangerous.
In combustion, toxic chemicals are sorbed by particulates and thus can be inhaled, so that the
emitted particles are doubly dangerous. It is argued by the proponents of incineration that fine
particulates constitute only a very small fraction by weight of total particulate emissions, so
that the danger is very small. However, that argument is false, because it is not the size of
particle that is important, but the number of particles and their surface area (because toxic
chemicals are sorbed on t o the surface). Atkinson et al [35] have very recently shown that
mortality and hospital admissions correlated with numbers of atmospheric particles, but not
with their weight. For a given weight concentration of particles, there are more of them if
they are smaller; Livingston [37] has estimated that one pound of very fine particles emitted
from an incinerator will consist of 140 quadrillion (1,000,000,000,000,000) particles. Another
way of appreciating this is to note that about 26,500 PM 2.5 particles would fit on the dot of
the letter i in normal print; for PM 0.001 particles the figure would be 160 billion. Also, for a
given weight concentration, smaller particles have greater surface area. A given weight of
PM 0.1 will have 100 times the surface area of the same weight of PM 10 ; in fact, in terms of
surface area, fine particulates constitute almost 50% of the total particulate emissions [38]. A
further problem is that, unlike larger particulates, fine particulates cannot be filtered
efficiently, and so are released into the atmosphere.
No information has been provided by Peel about the effects of particulate emissions from the
proposed biomass plant on “sensitive receptors”.
The U.K. already suffers from poor air quality, and the European Union is instigating
proceedings against the British Government, particularly in respect of levels of particulate
emissions [39, 40]. The north of England has worse pollution than does the south, as a recent
study published in the British Medical Journal showed [41]. Ellesmere Port had the highest
infant death rate in the U.K. in 2006, and Runcorn has one of the highest death rates in
England from lung cancer. It is unjustifiable to inflict yet another polluting industrial
plant on us. This is discussed in more detail in Section 8 (Environmental Justice) below.
10
There is a vast literature concerning the health effects of airborne particulate matter [42]. It is
now established beyond reasonable doubt that particulate air pollution can cause
cardiovascular morbidity and mortality [43], cardiopulmonary mortality [44], and respiratory,
immunological, haematological, neurological and reproductive/developmental problems [45],
sometimes with long time-lags between exposure and health effects. Pope et al [46] found
that each 10 µg/m3 increase in fine particulate levels was associated with a 4%, 6% and 8%
increased risk of all-cause, cardiopulmonary, and lung cancer mortality respectively. There is
particular concern about the effects of particulate pollution on infants. Woodruff et al [47]
found increases in infant deaths from respiratory causes with a 10 µg/m3 increase in PM 2.5 s.
Pino et al [48] found that a 10 µg/m3 increase in PM 2.5 s was related to a 5% increase in the
risk for wheezing bronchitis.
Still smaller particles (≤ 0.1 µm (100 nm) diameter) are termed nanoparticles. They are able
not only to penetrate most deeply into the lung, but are capable of being taken up
systemically, entering cells, disrupting cell signalling and other processes [49, 50, 51].
Howard [38] cited U.S. E.P.A. figures showing that for typical particulate incinerator
emissions 48.8% of the surface area is provided by particles of < 0.7 µm diameter. The
significance of this for toxicity is that toxic chemicals such as dioxins and heavy metals can
be sorbed on to the surfaces of particulate matter and taken into sensitive areas of the body.
Howard [38] also quoted figures from Onyx showing that baghouse filter collection
efficiency was 95-99% for PM 10 s, 65-70% for PM 2.5 s, and only 5-30% for particles smaller
than 2.5 µm, even before the filters become coated with lime and activated carbon. Brown et
al [52] have pointed out that long-term exposure to even low concentrations of fine particles
may be associated with reduced life expectancy.
Cormier et al [53] have reviewed the evidence for potential health impacts of incinerator
particulate emissions. They posed a series of questions that require answers:
How are combustion-generated fine PM and ultrafine PM formed?
How do their chemical properties differ from larger PM?
What is the nature of association of chemicals with these particles?
How is the chemical and biological reactivity of these chemicals changed by association with
the particles?
What is the role of PM-associated persistent free radicals in the environmental impacts of fine
and ultrafine PM?
What is the role of PM on cell/organ functioning at initial sites of exposure?
What is the bioavailability of these particles to other tissues?
How are these particles translocated to these secondary sites, and do their chemical properties
change en route?
How does acute/chronic exposure lead to adverse organ pathophysiology? Is developmental
timing of exposure important?
What effect does exposure have on predisposing to disease states or on disease progression?
Most importantly, what are the specific cellular and molecular mechanisms associated with
airborne exposures?
It is clear from the above that medical science has only very recently started to recognise the
serious problems that particulate emissions can cause, and it will be many years before the
answers to the questions posed above are available. Meanwhile it is essential that particulate
emissions, especially those produced in conjunction with toxic chemicals, are reduced.
Because Peel will no d oubt argue that the proposed Ince biomass plant would emit low
quantities of particulates, it should be stressed here that cumulative effects on health due to
continual exposure to environmental pollutants can be very serious even at levels below the
11
national ambient air quality standards of America [54]. Incineration, whether of waste or
biomass, is therefore a dangerous option for waste treatment.
(c) Ash. Combustion of wood leaves a solid residue of ash, which can contain metals such as
chromium and lead, and dioxins. This means the ash must be treated as hazardous waste, and
not (as sometimes happens) used as fertiliser, ingredients in cement, and road base.
The Health Protection Agency (HPA) issued a report in 2009 [55], in which they stated:
“While it is not possible to rule out adverse health effects from modern, well regulated
municipal waste incinerators with complete certainty, any potential damage to the health of
those living close-by is likely to be very small, if detectable”. In my professional view, as
an environmental scientist, that statement is unjustified. The HPA report contains many
errors and unjustified assumptions (for example, it gives an incorrect and very misleading
discussion of particulates), and refers to only eight peer-reviewed published studies. Those
studies were all flawed, for one very simple reason – they failed to take account of wind
direction. It is clear that any effects of pollutants emitted to atmosphere will be greater
downwind, and yet not one of the studies cited, or indeed any of the numerous other
published studies of health effects near incinerators, has taken that into account.
Three things follow from that:
(i) Any adverse effects found in those non-directional studies must actually be quite
significant. Miyake et al [30] found increased prevalence of wheeze, headache, stomach ache
and fatigue in children living close to incinerators, after allowing for confounding factors
such as socioeconomic status; Cordier et al [26] found an increased risk of birth defects in
children born to mothers living near incinerators; Viel et al [56] found increased incidences
of soft tissue sarcoma and non-Hodgkin’s lymphoma in the vicinity of a waste incinerator,
which was stated as unlikely to be due to confounding factors.
(ii) Government agencies are either less than meticulous in their examination of published
studies, or are turning a blind eye to flaws and irregularities in those studies.
(iii) Peer review does not guarantee that a published paper is scientifically valid.
The effect of wind direction is 5demonstrated using the wind-rose for Ellesmere Port:
345
15
355 900
N
335
25
325
1999
35
800
1998
45
700
600
315
1997
2000
55
2001
500
305
65
400
295
75
300
200
285
275
85
100
E
0
95
W
265
105
115
255
245
125
235
135
225
145
215
155
205
195
S
165
175
Figure 3. Wind-rose data for Ellesmere Port for the years 1997-2001 (kindly supplied by the
former Ellesmere Port and Neston Borough Council)
12
It can be seen that the majority of directions receive little or no wind. The main winds come
from the WNW and just E of S, each encompassing a sector of about 60°, which together
make ⅓ of the total 360°. Suppose, for the sake of argument, that cancer risk in the two main
sectors is twice that of the remaining 240° because of incinerator emissions. If wind direction
is ignored, then the cancer risk for all directions is (2 x ⅓) + (1 x ⅔) = 1.33, w hich is
much less than 2, and is probably statistically not significantly different from unity, bearing
in mind the uncertainties in epidemiological data.
So far as I am aware, the only work on health effects around incinerators that has taken
account of wind direction is that of Michael Ryan, and is unpublished. However, it does not
attempt to take account of confounding effects. An example is shown in Figure 4.
Figure 4. Infant (under 1 year) deaths near the Edmonton incinerator for 2003-2005, showing
the effect of wind direction (from Michael Ryan)
Clearly, what is still required, and to date has never been done, is an extensive study that
takes account of both confounding effects and wind direction. Until such a study is done, we
need to take a precautionary approach, which will preclude the siting of incinerators,
including biomass incinerators, within, say, 10 km of residential areas.
5. Regulatory controls
Peel’s proposed biomass incinerator is a 20 MW plant. It should be noted that whilst stringent
emissions standards apply to units over 20 MW, below that there are no regulations that apply
across the U.K. It is not known whether Peel’s plant would fall into the former or latter
category.
It should also be noted that a company called Prenergy, which obtained planning permission
for a very large (350 MW) biomass incinerator in Port Talbot, south Wales in 2009, have now
applied for permission to increase their emission limits for nitrous oxide from 20 mg/m3 to 40
13
mg/m3, for sulphur dioxide from 10 m g/m3 to 50 m g/m3, and for hydrogen chloride
(hydrochloric acid) from 7 mg/m3 to 10 mg/m3, and the Environment Agency Wales has said
that it is likely to approve these changes. This is a standard planning dodge, and steps
should be taken to ensure that Peel not do this.
6. Visual amenity
The proposed biomass plant would replace the bio-ethanol plant. However, the stack height
would increase from 30 m etres to 85 m etres, a very significant increase. In addition, the
height of the main building would increase from 28 metres to 42 metres – again, a significant
increase. It is clear that the biomass plant would have a major visual impact. A further point
is that the elevation plans provided by Peel show a height of 49 m etres, not 42 m etres as
stated in the Environmental Statement and the Planning Statement. Which height is correct?
I consider the statement that the proposed Frodsham Marsh wind farm development will
reduce the magnitude of change introduced by the biomass proposal to be quite unjustified.
Firstly, the wind farm development has not been granted planning permission, and may be
refused. Secondly, in their planning application for the wind farm, Peel argued that it would
not be obtrusive and would not damage the openness of Frodsham Marsh. I do not agree with
that argument, but Peel cannot have it both ways.
7. Perception of risk
In their “Rapid Health Impact Assessment of the proposed Ince Resource Recovery Park”,
Western Cheshire Primary Care Trust [57] drew attention to the importance of risk perception
by the public relating to the siting and presence of facilities that might be construed as posing
a threat to health or amenity. Starr and Whipple [58] developed a quantitative approach to
perceived risk assessment, based on the assumption that society achieves, by trial and error, a
reasonable balance between risk and benefit. They drew the following conclusions:
1. the acceptability of risk is roughly proportional to the real and perceived benefits;
2. voluntary risks are some 1000 times more acceptable than are involuntary risks;
3. the tolerable level of risk is inversely related to the number of involved persons.
Since there appear to be few, if any, benefits from burning biomass, since the risk posed by
the biomass plant would be involuntary, and since there are thousands of local residents in
Ellesmere Port, Ince, Elton, Thornton-le-Moors, Helsby, Frodsham and Runcorn who would
be affected by the plant’s emissions, it is clear that in this case there would be a far from
reasonable balance between risk and benefit.
According to the Department of Health [59], risks are generally more worrying and less
acceptable if perceived:
1.
2.
3.
4.
5.
6.
7.
to be involuntary (e.g. exposure to pollution) rather than voluntary (e.g. dangerous
sports or smoking);
as inequitably distributed (some benefit whilst others suffer the consequences);
as inescapable by taking personal precautions;
to arise from an unfamiliar or novel source;
to result from man-made rather than natural sources;
to cause hidden and irreversible damage (e.g. through onset of illness many years
after exposure);
to pose some particular danger to small children or pregnant women or more
generally to future generations;
14
8.
9.
10.
11.
to threaten a form of death (or illness/injury) arousing particular dread;
to damage identifiable rather than anonymous victims;
to be poorly understood by science;
as subject to contradictory statements from responsible sources.
All of these perceptions apply to the proposed Ince biomass incinerator, and are especially
relevant at present with everyone’s attention focussed on t he potential nuclear reactor
catastrophe in Japan. Whilst there is no c omparison in terms of magnitude of risk, the
Japanese situation has caused all governments to re-assess potential risks, and has lessened
the faith of the public in the ability of regulators and companies to control these, and of the
authorities to tell them the truth. This can only add to the concerns over the proposed Ince
biomass plant.
Lima [60] has pointed out that “incinerators represent a solution to urban waste problems in
which most of the beneficiaries (those who produce the waste) are not exposed to the risks
and to the inconveniences of the facility; on the (other hand), those who live near the site face
all the problems during construction and (during) normal functioning of the station. They
have to deal with the unpleasant changes to their environment and the uncertainty about the
health consequences of the facility, and they consider this situation to be unfair”.
Gregory et al [61] have pointed out that incineration is a hazard with characteristics such as
dread consequences and involuntary exposure, its impacts are perceived to be inequitably
distributed, and its effects are unbounded in the sense that their magnitude or persistence over
time is not well known. Lima [60], in a 5-year assessment of the effect of risk perception on
the mental health of people living near an incinerator, found that: (i) risk perception is more
acute for residents living closer to the site, who also have a less favourable attitude; (ii) there
is an habituation effect for those living closer to the incinerator; (iii) psychological symptoms
are associated with socio-economic variables (sex and education), but also with
environmental annoyance; (iv) for those living close to the site, risk perception and the
interaction between risk perception and environmental annoyance significantly increase the
prediction of psychological symptoms such as stress, anxiety and depression. This was after
the confounding effects of other environmental stressors such as noise and traffic had been
allowed for. Lima’s results confirm the Lazarus and Folkman [62] proposal that the health
consequences of an environmental stressor depend on t he appraisal of the threat and of the
personal resources to deal with it. That is, risk perception per se can modify the quality of life
of those living under suspicion of objective risk [63]. Stress effects can persist for many
years, as was shown by Matthies et al [64] in a study of people living in an area with
contaminated soil.
Wandersman and Hallman [65] have pointed out that: “Risk perceptions that do not match
scientific estimates of risk are not necessarily irrational. For example, if one believes that
regulators cannot be trusted, experts are not well trained, and accidents can easily be caused
by human error, then it is rational to view that risk as unacceptable”. Maynard [66], of the
U.K. Government’s Department of Health, pointed out that “a risk can only be described as
acceptable if the public regard it as acceptable. The role of the scientist is to provide the
public with the best possible basis for reaching their decision; it is not for scientists to decide
whether a risk is acceptable though they will have their own views on this. It is clear that the
acceptability of a risk will depend on t he confidence that the public has in the process that
leads to this risk assessment”. Maynard summarised the public’s attitude to risk as: worry =
risk x fear.
15
In summary, there is considerable evidence that the presence or proposed presence of
facilities that might be construed as posing a threat to health or amenity can cause mental
health problems in populations. I believe that perception of risk from the proposed Ince
biomass plant is likely to lead to an increase in stress-related disorders in the local population,
and thereby place an additional burden on local medical services.
8. Environmental justice
A factor related to risk perception is environmental justice, to which the Western Cheshire
Primary Care Trust [57] drew attention. Their report pointed out that: “There is a perception
amongst the local communities that the area is already overdeveloped and the community is
being ‘dumped on a gain’ by industry”. This area indeed has far more than its fair share of
heavy and polluting industry: the huge Essar (formerly Shell) oil refinery (including a waste
incinerator), GrowHow UK, Quinn Glass, Air Products, Veolia (ES) Cleanaway (with high
temperature incinerator), Innospec, Tradebe Waste Management, Greif UK, Electrical Oil
Services (with commercial waste oil recovery), IneosChlor, and IneosFluor. In addition, there
is planning permission for two huge waste incinerators, one of 600,000 tonne/year capacity at
the Ince Resource Recovery Park, and one at Ineos, Runcorn of 850,000 tonne/year capacity.
The Scottish Executive [67] has stated that, for environmental justice to be done, “deprived
communities which may be more vulnerable to the pressures of poor environmental
conditions should not bear a disproportionate burden of negative environmental impacts”. It
should be noted here that both Ellesmere Port and Runcorn have areas of significant
deprivation; the former has a very high infant mortality rate, and the latter has a v ery high
lung cancer mortality rate.
A salutary example of the way in which our area has long suffered environmental injustice
was reported by Woods [68]. In discussing the expansion of Elton in the 1970s, he quoted a
developer who submitted a planning application for residential development in Elton in 1970:
“The land around the holding is mostly developed for industrial or residential use. The
holding is not large enough to be worked so as to provide a living and the factory effluence in
the air does not encourage good grass to grow. In consequence, we feel that residential
development is the only logical use to which this land can be put”. The water authority were
consulted, and raised no objection, but the Public Health Inspector and the Alkali Inspector
were not asked for their views. Cheshire County Council refused the application on G reen
Belt grounds.
9. Summary
I have shown that there are many reasons for the proposed Ince biomass plant to be refused
planning permission. These include lack of need, lack of carbon neutrality, the waste
hierarchy, toxic emissions, regulatory controls, visual amenity, perception of risk, and
environmental justice. I ask that Cheshire West and Chester Council refuse planning
permission for this plant.
16
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20
Environmental Health
BioMed Central
Open Access
Review
Systematic review of epidemiological studies on health effects
associated with management of solid waste
Daniela Porta1, Simona Milani1, Antonio I Lazzarino1,2, Carlo A Perucci1 and
Francesco Forastiere*1
Address: 1Department of Epidemiology, Regional Health Service Lazio Region, Rome, Italy and 2Division of Epidemiology, Public Health and
Primary Care, Imperial College, London, UK
Email: Daniela Porta - [email protected]; Simona Milani - [email protected]; Antonio I Lazzarino - [email protected];
Carlo A Perucci - [email protected]; Francesco Forastiere* - [email protected]
* Corresponding author
Published: 23 December 2009
Environmental Health 2009, 8:60
doi:10.1186/1476-069X-8-60
Received: 4 May 2009
Accepted: 23 December 2009
This article is available from: http://www.ehjournal.net/content/8/1/60
© 2009 Porta et al; licensee BioMed Central Ltd.
This is an Open Access article distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/2.0),
which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.
Abstract
Background: Management of solid waste (mainly landfills and incineration) releases a number of toxic
substances, most in small quantities and at extremely low levels. Because of the wide range of pollutants, the
different pathways of exposure, long-term low-level exposure, and the potential for synergism among the
pollutants, concerns remain about potential health effects but there are many uncertainties involved in the
assessment. Our aim was to systematically review the available epidemiological literature on the health effects in
the vicinity of landfills and incinerators and among workers at waste processing plants to derive usable excess risk
estimates for health impact assessment.
Methods: We examined the published, peer-reviewed literature addressing health effects of waste management
between 1983 and 2008. For each paper, we examined the study design and assessed potential biases in the effect
estimates. We evaluated the overall evidence and graded the associated uncertainties.
Results: In most cases the overall evidence was inadequate to establish a relationship between a specific waste
process and health effects; the evidence from occupational studies was not sufficient to make an overall
assessment. For community studies, at least for some processes, there was limited evidence of a causal
relationship and a few studies were selected for a quantitative evaluation. In particular, for populations living
within two kilometres of landfills there was limited evidence of congenital anomalies and low birth weight with
excess risk of 2 percent and 6 percent, respectively. The excess risk tended to be higher when sites dealing with
toxic wastes were considered. For populations living within three kilometres of old incinerators, there was limited
evidence of an increased risk of cancer, with an estimated excess risk of 3.5 percent. The confidence in the
evaluation and in the estimated excess risk tended to be higher for specific cancer forms such as non-Hodgkin's
lymphoma and soft tissue sarcoma than for other cancers.
Conclusions: The studies we have reviewed suffer from many limitations due to poor exposure assessment,
ecological level of analysis, and lack of information on relevant confounders. With a moderate level confidence,
however, we have derived some effect estimates that could be used for health impact assessment of old landfill
and incineration plants. The uncertainties surrounding these numbers should be considered carefully when health
effects are estimated. It is clear that future research into the health risks of waste management needs to overcome
current limitations.
Page 1 of 14
(page number not for citation purposes)
Introduction
"Waste management", that is the generation, collection,
processing, transport, and disposal of solid waste is
important for both environmental reasons and public
health. There are a number of different options available
for the management and treatment of waste including
minimisation, recycling, composting, energy recovery and
disposal. At present, an increasing amount of the
resources contained in waste is recycled, but a large portion is incinerated or permanently lost in landfills. The
various methods of waste management release a number
of substances, most in small quantities and at extremely
low levels. However, concerns remain about potential
health effects associated with the main waste management technologies and there are many uncertainties
involved in the assessment of health effects.
Several studies of the possible health effects on populations living in proximity of landfills and incinerators have
been published and well-conducted reviews are available
[1-4]. Both landfills and incinerators have been associated
with some reproductive and cancer outcomes. However,
the reviews indicate the weakness of the results of the
available studies due to design issues, mainly related to a
lack of exposure information, use of indirect surrogate
measures, such as the distance from the source, and lack
of control for potential confounders. As a result, there is
great controversy over the possible health effects of waste
management on the public due to differences in risk communication, risk perception and the conflicting interests
of various stakeholders. Therefore, there is the need for an
appropriate risk assessment that informs both policy makers and the public with the information currently available on the health risks associated with different waste
management technologies. Of course, the current uncertainties should be taken into account.
Within the EU-funded INTARESE project [5], we aimed to
assess potential exposures and health effects arising from
solid wastes, from generation to disposal, or treatment. A
key part in the health impact assessment was selecting or
developing a suitable set of relative risks that link individual exposures with specific health endpoints. In this
paper, we systematically reviewed the available epidemiological literature on health effects in the vicinity of landfills and incinerators and among workers at waste
processing plants to derive usable excess risk estimates for
health impact assessment. The degree of uncertainty associated with these estimates was considered.
Methods
We considered epidemiological studies conducted on the
general population with potential exposures from collecting, recycling, composting, incinerating, and landfilling
solid waste. We also considered studies of employees of
http://www.ehjournal.net/content/8/1/60
waste management plants as they may be exposed to the
same potential hazards as the community residents, even
if the intensity and duration of the exposure may differ.
However, to limit our scope, we did not consider studies
on biomarkers of exposure and health effects.
Relevant papers were found through computerized literature searches of MEDLINE and PubMed Databases from
1/1/1983 through 31/12/2008, using the MeSH terms
"waste management" and "waste products" and the subheading "adverse effects". We identified 144 papers with
this method. We also conducted a free search with several
combinations of relevant key words (waste incinerator or
landfill or composting or recycling) and (cancer or birth
outcome or health effects), and 285 papers were identified. In addition, articles were traced through references
listed in previous reviews [1-3,6-9], and in publications of
the UK Department for Environment, Food and Rural
Affairs [10]. Finally, we used information from two recent
reviews of epidemiological studies on populations with
potential exposures from toxic and hazardous wastes for
reproductive [4], and cancer [11] outcomes, respectively.
The eligibility of all papers was evaluated independently
by three observers, and disagreements were resolved by
discussion. As indicated, studies on sewage treatment and
on biological monitoring were not included. We also
excluded articles in languages other than English, not
journal articles, and six studies [12-17] conducted at the
municipal level (usually small towns) where it was not
possible to evaluate the extent of the population potentially involved and the possibility of exposure misclassification was high.
Papers were grouped according to the following criteria:
• waste management technologies: recycling, composting,
incinerating, landfilling (considering controlled disposal
of waste land and toxic or hazardous sites);
• health outcomes: cancers (stomach, colorectal, liver, larynx and lung cancer, soft tissue sarcoma, kidney and bladder cancer, non-Hodgkin's lymphoma, childhood
cancer), birth outcomes (congenital malformations, low
birth weight, multiple births, abnormal sex ratio of newborns), respiratory, skin and gastrointestinal symptoms or
diseases.
We have reported in the appropriate tables (in the online
additional files) for each paper: study design (e.g. geographical, cohort, cross-sectional, case-control study,
etc.), population characteristics (subjects, country, age,
sex), exposure measures (e.g. occupational exposure to
waste incinerator by-products, residence near a landfill,
etc.), and the main results (including control for major
Page 2 of 14
confounders) with respect to the quantification of the
health effects studied. For each study we have evaluated
the potential sources of uncertainty in the results due to
design issues. In particular, the possibility that selection
bias, information bias, or confounding could artificially
increase or decrease the relative risk estimate has been
noted in the tables using the plus/minus scale to indicate
that effect estimates are likely to be overestimated (or
underestimated) up to 20% (+/-), from 20 to 50% (++/--)
and more than 50% (+++/---). Uncertainties were graded
by two observers (SM and FF), who discussed the inconsistencies.
cial, etc, are variously applied in different countries and
time periods to designate non-household wastes. In earlier time periods definitions were even less clear and some
disposal sites may have switched categories (e.g. if they
used to take industrial waste they may now only take
municipal waste). Since two systematic reviews were
already available for toxic wastes [4,11], we did not replicate the literature search, but summarized the evidence
reported in the available reviews and tried to compare and
discuss the results with studies where mainly municipal
solid wastes were landfilled. The additional file 1 contain
several details of the studies reviewed.
After a description of the available studies, the overall
evaluation of the epidemiological evidence regarding the
process/disease association was made based on the IARC
(1999) criteria, and two categories were chosen, namely:
"Inadequate" when the available studies were of insufficient quality, consistency, or statistical power to determine the presence or absence of a causal association;
"Limited" when a positive association was observed
between exposure and disease for which a causal interpretation is considered to be credible, but chance, bias, or
confounding could not be ruled out with reasonable confidence. There were no instances where the category "sufficient" evidence could be used. Only when the specific
process/disease association was judged as limited (suggestive evidence but not sufficient to infer causality) we
decided to evaluate the strength of the association and to
measure appropriate relative risks. For this purpose, we
considered the set of studies providing the best evidence
and assigned an overall level of scientific confidence of
the specific effect estimate based on an arbitrary scale: very
high, high, moderate, low, very low. This evaluation was
made by three assessors (SM, DP, and FF).
Cancer
Russi et al. [11] carried out Medline searches of the peerreviewed English language medical literature covering the
period from January 1980 to June 2006 using the keywords "toxic sites" and "cancer", and identified articles
from published reviews. They included 19 articles which
fit the following selection criteria: 1) the study addressed
either cancer incidence or cancer mortality as an endpoint, 2) the study was carried out in a community or a set
of communities containing a known hazardous waste site;
3) the study had to address exposure from a specific waste
site, rather than from a contaminated water supply
resulted from multiple point sources. As the authors recognized, some of the location investigated included both
toxic wastes and municipal solid wastes as in the study
from Goldberg et al. [18] or Pukkala et al. [19]. There are
two investigations considered in this review that are
important to evaluate because of the originality of the
approach (cohort study, [19] and due to the large size
[20].
Results
A total of 49 papers were reviewed: 32 concerning health
effects in communities in proximity to waste sites, and 17
on employees of waste management sites. The majority of
community studies evaluated possible adverse health
effects in relation to incinerators and landfills. We found
little evidence on potential health problems resulting
from environmental or occupational exposures from
composting or recycling, and very little on storage/collection of solid waste. A description of the main findings follows.
Studies of communities near landfills
One of the main problems in dealing with studies on
landfill sites (an to some extent also for incinerators) is
the distinction between sites for municipal solid wastes
and sites for other wastes. The definition of different types
of waste is far from being standardised across the world.
The terms hazardous, special, toxic, industrial, commer-
In Finland, Pukkala et al. [19] studied whether the exposure to landfills caused cancer or other chronic diseases in
inhabitants of houses built on a former dumping area
containing industrial and household wastes. After adjusting for age and sex, an excess number of male cancer cases
were seen, especially for cancers of the pancreas and of the
skin. The relative risk slightly increased with the number
of years lived in the area. However, some uncertainties
were likely to affect the results of the study with regards to
the exposure assessment (-), outcome assessment (+) and
presence of residual confounding (-).
Jarup et al. [20] examined cancer risks in populations living within 2 km of 9,565 (from a total of 19,196) landfill
sites that were operational at some time from 1982 to
1997 in Great Britain. No excess risks of cancers of the
bladder and brain, hepato-biliary cancer or leukaemia
were found, after adjusting for age, sex, calendar year and
deprivation. The study was very large and had high power,
however misclassification of exposure could have
decreased the possibility of detecting an effect (--).
Page 3 of 14
Based on the findings and on the evaluation of the quality
of the studies, Russi et al. [11] concluded that epidemiological studies of populations living in the vicinity of a
toxic waste site have not produced evidence of adequate
quality to establish a casual link between toxic waste exposures and cancer risk. In our terms, the evidence may be
considered as "inadequate".
In addition to the articles reviewed by Russi et al. [11], we
reviewed the article by Michelozzi et al. [21], which investigated the mortality risk in a small area of Italy (Malagrotta, Rome) with multiple sources of air contamination
(a very large waste disposal site serving the entire city of
Rome, a waste incinerator plant, and an oil refinery
plant). Standardised Mortality Ratios (SMRs) were computed in bands of increasing distance from the plants, up
to a radius of 10 km. No association was found between
proximity to the sites and cancer of various organs, in particular liver, lung, and lymph haematopoietic cancer,
however, mortality from laryngeal cancer declined with
distance from the pollution sources, and a statistically significant trend remained after adjusting for a four-level
index of socio-economic status. The main uncertainty of
the study is related to the exposure assessment (--) since
only distance was considered thus decreasing the possibility of detecting an effect. There are also uncertainties in
using mortality to estimate cancer incidence in proximity
to a suspected source of pollution (+). On the other hand,
even though the authors did adjust for an area-based
index of deprivation, residual confounding (+) from socioeconomic status was likely.
In summary, there is inadequate evidence of an increased
risk of cancer for communities in proximity of landfills.
The three slightly positive studies from Goldberg et al.
[18], Pukkala et al. [19] and Michelozzi et al. [21] are not
consistent.
Birth defects and reproductive disorders
Saunders [4] reviewed 29 papers examining the relationship between residential proximity to landfill sites and the
risk of an adverse birth outcome. The review included
either studies on municipal waste or on hazardous waste.
Eighteen papers reported some significant association
between adverse reproductive outcome and residence
near a landfill site. Two of the strongest papers conducted
on hazardous waste landfill sites in Europe (EUROHAZCON) found similarly moderate but significant associations between residential proximity (within 3 km) to
hazardous waste sites and both chromosomal [22] (Odds
Ratio, OR: 1.41, 95%CI: 1.00-1.99) and non-chromosomal [23] (OR: 1.33, 95%CI: 1.11-1.59) congenital
anomalies.
Included in the Saunders's review [4] is the national geographical comparison study on landfills in the UK by Elliott et al. [24]. This study investigated the risk of adverse
birth outcomes in populations living within two km of
9,565 landfill sites in Great Britain, operational at some
time between 1982 and 1997, compared with those living
further away (reference population). The sites included
774 sites for special (hazardous) waste, 7803 for non-special waste and 988 handling unknown waste; a two km
zone was defined around each site to detect the likely
limit of dispersion for landfill emissions, including 55%
of the national population. Among the 8.2 million live
births and 43,471 stillbirths, 124,597 congenital anomalies (including miscarriage) that were examined, there
were: neural tube defects, cardiovascular defects, abdominal wall defects, hypospadias and epispadias, surgical correction of gastroschisis and exomphalos; low and very low
birth weights were also found , defined as less than 2500
g and less than 1500 g, respectively. The main analysis,
conducted for all landfill sites during their operation and
after closure, found a small, but still statistically significant, increased risk of total and specific anomalies (OR:
1.01, 95%CI: 1.005-1.023) in populations living within 2
Km, and also an increased risk of low (OR: 1.05, 95%CI:
1.047-1.055) and very low birth weight (OR: 1.04,
95%CI: 1.03-1.05). Additional analyses were carried out
separately for sites handling special waste and non-special
waste, and in the period before and after opening, for the
5,260 landfills with available data. After adjusting for deprivation and other potential confounding variables (sex,
year of birth, administrative region), there was a small
increase in the relative risks for low and very low birth
weight and for all congenital anomalies, except for cardiovascular defects. The risks of all congenital anomalies
were higher for people living near special waste disposals
(OR: 1.07 CI95%:1.04-1.09) compared to non-special
waste disposals (OR: 1.02, CI95%:1.01-1.03). There was
no excess risk of stillbirth. On these bases, the author [4]
concluded that while most studies reporting a positive
association are of good quality, over half report no association with any adverse birth outcome and most of the latter are also well conducted. The review considered that the
evidence of an association of residence near a landfill with
adverse birth outcomes as unconvincing.
After the review by Saunders [4], we considered four additional studies examining reproductive effects of landfill
emissions.
Elliot et al. recently updated the previous study [25] in
order to evaluate whether geographical density of landfill
sites was related to congenital anomalies. The analysis was
restricted to 8804 sites operational at some time between
1982 and 1997. There were 607 sites handling special
(hazardous) waste and 8197 handling non-special or
Page 4 of 14
unknown waste type. The exposure assessment took into
account the overlap of the two km buffers around each
site, to define an index of exposure with four levels of
increasing landfill density. Several anomalies (hypospadias and epispadias, cardiovascular defects, neural tube
defects and abdominal wall defects) were evaluated. The
analysis was carried out separately for special and nonspecial waste sites and was adjusted for deprivation, presence or absence of a local congenital anomalies register
and maternal age. The study found a weak association
between intensity of hazardous sites and some congenital
anomalies (all, cardiovascular, hypospadia and epispadias).
retardation. The major limit of the study is the low specificity of the exposure definition.
The studies conducted in the United Kingdom suffer from
the same limitations, namely the possibility that misclassification of exposure could have decreased the relative
risk estimates to some extent (--); on the other hand, there
are several uncertainties related to the quality of reporting
and registration of congenital malformations. In the latter
case, a positive bias is more likely (++). For the recent
report by Elliott et al. [25], location uncertainties and differential data reliability regarding the sites, together with
the use of distance as the basis for exposure classification,
limit the interpretation of the findings (--).
Respiratory diseases
A study conducted by Pukkala et al. [19] in Finland evaluated prevalence of asthma in relation to residence in
houses built on a former dumping area containing industrial and household wastes. Prevalence of asthma was significantly higher in the dump cohort than in the reference
cohort (living nearby but outside the landfill site). Unfortunately, this study has not been replicated and the overall
evidence may be considered inadequate.
In Denmark, Kloppenborg et al. [26] marked the geographical location of 48 landfills and used maternal residence as the exposure indicator in a study of congenital
malformations. The authors found no association
between landfill location and all congenital anomalies or
of the nervous system, and a small excess risk for congenital anomalies of the cardiovascular system. Potential confounding from socioeconomic status is the major
limitation of this study (+++).
Jarup et al. [27] studied the risk of Down's syndrome in
the population living near 6829 landfills in England and
Wales. People were considered exposed if they lived in a
two-km zone around each site, people beyond this zone
were the reference group. A two-year lag period between
potential exposure of the mother and her giving birth to a
Down's syndrome child was allowed. The analysis was
adjusted for maternal age, urban-rural status and deprivation index. No statistically significant excess risk was
found in the exposed populations, regardless of waste
type.
Finally, Gilbreath et al. [28] studied births in 197 Native
Alaskan villages containing open dumpsites with hazardous waste, scoring the exposure into high, intermediate
and low hazard level on the basis of maternal residence.
The authors found an association between higher levels of
hazard and low birth weight and intrauterine growth
In summary, an increased risk of congenital malformations and of low birth weight has been reported from
studies conducted in the UK. When compared with the
results from studies conducted in proximity of hazardous
waste sites, studies in proximity of non-toxic waste landfills provide lower effect estimates. The main uncertainty
of these studies is the completeness of data on birth
defects, the use of distance from the sites for exposure classification, and the classification as toxic and non-toxic
waste sites.
Studies of landfills workers
Only one study on landfill workers was reviewed. Gelberg
et al. [29] conducted a cross-sectional study to examine
acute health effects among employees working for the
New York City Department of Sanitation, focusing on
Fresh Kills landfill employees. Telephone interviews conducted with 238 on-site and 262 off-site male employees
asked about potential exposures both at home and work,
health symptoms for the previous six months, and other
information (social and recreational habits, socio-economic status). Landfill workers reported a significantly
higher prevalence of work-related respiratory, dermatological, neurologic and hearing problems than controls.
Respiratory and dermatologic symptoms were not associated with any specific occupational title or task, other than
working at the landfill, and the association remained,
even after controlling for smoking status.
Studies of communities living near incinerators
Twenty-one epidemiologic studies conducted on residents of communities with solid waste incinerators have
been reviewed and their characteristics are listed in the
additional file 2.
Cancer
Eleven studies have been reviewed on cancer risk in relation with incinerators, usually old plants with high polluting characteristics. The studies are reported below by
country.
Page 5 of 14
In the United Kingdom, Elliott et al. [30] investigated cancer incidence between 1974 and 1987 among over 14 million people living near 72 solid waste incinerator plants.
Data on cancer incidence among the residents, obtained
from the national cancer registration programme, were
compared with national cancer rates, and numbers of
observed and expected cases were calculated after stratifying for deprivation, based on the 1981 census. Observedexpected ratios were tested for decline in risk up to 7.5 km
away. The study was conducted in two stages: the first
involved a stratified random sample of 20 incinerators
and, based on the findings, a number of cancers were then
further studied around the remaining 52 incinerators (second stage). Over the two stages of the study there was a
statistically significant (p < 0.05) decline in risk with distance from incinerators for all cancers, stomach, colorectal, liver and lung cancer. The use of distance as the
exposure variable in this study could have led to some
degree of misclassification (--). On the other hand, the
same authors observed that residual confounding (+) as
well as misdiagnosis (+) might have increased the risk
estimates. When further analyses were made, including a
histological review of liver cancer cases [31], the risk estimates were lower (0.53-0.78 excess cases per 105 per year
within 1 km, instead of 0.95 excess cases per 105 as previously estimated).
Using data on municipal solid waste incinerators from the
initial study by Elliott et al. [30], Knox [32] examined a
possible association between childhood cancers and
industrial emissions, including those from incinerators.
From a database of 22,458 cancer deaths that occurred in
children before their 16th birthday between 1953 and
1980, he extracted 9,224 cases known to have moved at
least 0.1 km in their life time, and using a newly developed technique of analysis, he compared distances from
the suspected sources to the birth addresses and to the
death addresses. The childhood-cancer/leukaemia data
showed highly significant excesses of moves away from
birthplaces close to municipal incinerators, but the specific effects of the municipal incinerators could not be separated clearly from those of nearby industrial sources of
combustion. Misclassification of exposure is the main
limit of this paper (--).
In France, Viel et al. [33] detected a cluster of patients with
non-Hodgkin's lymphoma (NHL) and soft tissue sarcoma
around a French municipal solid waste incinerator with
high dioxin emissions. To better explore the environmental origin of the cluster suggested by these findings, Floret
et al. [34] carried out a population-based case-control
study in the same area, comparing 222 incident cases of
NHL diagnosed between 1980 and 1995 and controls randomly selected from the 1990 census. The risk of developing lymphomas was 2.3 times higher among individuals
living in the area with the highest dioxin concentration
than among those in the area with the lowest concentration. Given that a model was used to attribute exposure to
cases and controls, a random misclassification could have
reduced the effect estimates (--). Based of these results, a
nationwide study on NHL was conducted [35]. A total of
13 incinerators in France were investigated and dispersion
modelling was used to estimate ground-level dioxin concentration. Information about the exposure levels and
potential confounders was available at the census block
level. A positive association between dioxin level and
NHL was found with a stronger effect among females.
Although the study represents an improvement regarding
exposure assessment compared to investigations based on
distance from the source, it should be noted that the analysis was conducted at the census block level and the possibility of misclassification of the exposure (-) as well as of
residual confounding from socioeconomic status (+)
remains.
Viel et al. [36] have recently reported the findings from a
case-control study on breast cancer. There was no association or even a negative association between exposure to
dioxin and breast cancer in women younger or older than
60 years, respectively, living near a French municipal solid
waste incinerator with high exposure to dioxin. Design
issues and residual confounding from age and other factors (---) limit the interpretations of the study.
In Italy, Biggeri et al. [37] conducted a case-control study
in Trieste to investigate the relationship between multiple
sources of environmental pollution and lung cancer.
Based on distance from the sources, spatial models were
used to evaluate the risk gradients and the directional
effects separately for each source, after adjusting for age,
smoking habits, likelihood of exposure to occupational
carcinogens, and levels of air particulate. The results
showed that the risk of lung cancer was inversely related
to the distance from the incinerator, with a high excess relative risk very near the source and a very steep decrease
moving away from it. The main problem of the study is
the difficulty to separate the effects of other sources of pollution based on distance, and the possibility of potential
confounding from other sources remains (++). An excess
risk of lung cancer was also found in females living in two
areas of the province of La Spezia (Italy) exposed to environmental pollution emitted by multiple sources, including an industrial waste incinerator [38]. Again in this
study the limited exposure assessment could have
decreased the risk estimates (--), but positive confounding
from other sources is very likely.
A case-control study by Comba et al. [39] showed a significant increase in risk of soft tissue sarcomas associated
with residence within two km of an industrial waste incin-
Page 6 of 14
erator in the city of Mantua, with a rapid decrease in risk
at greater distances. There is a slight likelihood that
increased attention to the diagnosis for this form of cancer
in the vicinity of the plant could have introduced a small
bias (+) in the risk estimate. Another case-control study,
carried out in the province of Venice by Zambon et al. [40]
analyzed the association between soft-tissue sarcoma and
exposure to dioxin in a large area with 10 municipal solid
waste incinerators. The authors found a statistically significant increase in the risk of sarcoma in relation to both the
level and the length of environmental modelled exposure
to dioxin-like substances. The results were more significant for women than for men.
In summary, although several uncertainties limit the overall interpretation of the findings, there is limited evidence
that people living in proximity of an incinerator have
increased risk of all cancers, stomach, colon, liver, lung
cancers based on the studies of Elliott et al. [30]. Specific
studies on incinerators in France and in Italy suggest an
increased risk for non-Hodgkin's lymphoma, and soft-tissue sarcoma.
Birth defects and reproductive disorders
Six studies examined reproductive effects of incinerator
emissions (see additional file 2).
Jansson et al. [41] analysed whether the incidence of cleft
lip and palate in Sweden increased since operation of a
refuse incineration plant began. The results of this register
study, based on information from the central register of
malformations and the medical birth register, did not
demonstrate an increased risk.
A study by Lloyd et al. [42] examined the incidence of
twin births between 1975 and 1983 in two areas near a
chemical and a municipal waste incinerator in Scotland:
after adjusting for maternal age, an increased frequency of
twinning in areas exposed to air pollution from incinerators was seen. In the same study areas, Williams et al. [43]
investigated gender ratios, at various levels of geographical detail and using three-dimensional mapping techniques: analyses in the residential areas at risk from
airborne pollution from incinerators showed locations
with statistically significant excesses of female births.
To investigate the risk of stillbirth, neonatal death, and
lethal congenital anomaly among infants of mothers living close to incinerators (and crematoriums), Dummer et
al. [44] conducted a geographical study in Cumbria (Great
Britain). After adjusting for social class, year of birth, birth
order, and multiple births, there was an increased risk of
lethal congenital anomaly, in particular spina bifida and
heart defects.
Subsequently, Cordier et al. [45] studied communities
with fewer than 50,000 inhabitants surrounding the 70
incinerators that operated for at least one year from 1988
to 1997 in France. Each exposed community was assigned
an exposure index based on a Gaussian plume model,
estimating concentrations of pollutants per number of
years the plant had operated. The results were adjusted for
year of birth, maternal age, department of birth, population density, average family income, and when available,
local road traffic. The rate of congenital anomalies was not
significantly higher in exposed compared with unexposed
communities; only some subgroups of congenital anomalies, specifically facial cleft and renal dysplasia, were
more frequent in the exposed communities.
Tango et al. [46] investigated the association of adverse
reproductive outcomes with mothers living within 10 km
of 63 municipal solid waste incinerators with high dioxin
emission levels (above 80 ng international toxic equivalents TEQ/m3) in Japan. To calculate the expected number
of cases, national rates based on all live births, fetal deaths
and infant deaths occurred in the study area during 19971998 were used and stratified by potential confounding
factors available from the corresponding vital statistics
records: maternal age, gestational age, birth weight, total
previous deliveries, past experience of fetal deaths, and
type of paternal occupation. None of the reproductive
outcomes studied showed statistically significant excess
within two km of the incinerators, but a statistically significant decline in risk with distance from the incinerators
was found for infant deaths and for infant deaths with
congenital anomalies, probably due to dioxin emissions
from the plants.
In sum, there are multiple reports of increased risk of congenital malformations among people living close to incinerators but there are no consistencies between the
investigated outcomes. The overall evidence may be considered as limited. The study by Cordier et al. [45] provides the basis for risk quantifications at least for facial
cleft and renal dysplasia. Quantification for other reproductive disorders is more difficult.
Respiratory and skin diseases or symptoms
Four studies examined respiratory and/or dermatologic
effects of incinerator emissions (see additional file 2).
Hsiue et al. [47] evaluated the effect of long-term air pollution resulting from wire reclamation incineration on
respiratory health in children. 382 primary school children who resided in one control and three polluted areas
in Taiwan were chosen for this study. The results revealed
a decrement in pulmonary function (including forced
vital capacity and forced expiratory volume in one second) of those residents in the vicinity of incineration sites.
Page 7 of 14
Shy et al. [48] studied the residents of three communities
having, respectively, a biomedical and a municipal incinerator, and a liquid hazardous waste-burning industrial
furnace, and then compared results with three matchedcomparison communities. After adjustment for several
confounders (age, sex, race, education, respiratory disease
risk factors), no consistent differences in the prevalence of
chronic or acute respiratory symptoms resulted between
incinerator and comparison communities. Additionally,
no changes in pulmonary function between subjects of an
incinerator community and those of its comparison community resulted from the study by Lee et al. [49], based on
a longitudinal component from the Health and Clean Air
study by Shy et al. [48].
Miyake et al. [50] examined the relationship between the
prevalence of allergic disorders and general symptoms in
Japanese children and the distance of schools from incineration plants, measured using geographical information
systems. After adjusting for grade, socio-economic status
and access to health care per municipality, schools closer
to the nearest municipal waste incineration plant were
associated with an increased prevalence of wheeze and
headache; there was no evident relationship between the
distance of schools from such plants and the prevalence of
atopic dermatitis. The main factors that may have affected
the relative risk estimates in this study could be reporting
bias (++) and residual confounding from socioeconomic
status (++).
In sum, although the intensive study conducted by Shy et
al. [48] did not show respiratory effects, there are some
indications of an increased risk of respiratory diseases,
especially in children. However, the uncertainty related to
outcome assessment and residual confounding is very
high and the overall evidence may be considered inadequate.
Bresnitz et al. [52] studied 89 of 105 male incinerator
workers in Philadelphia, employed at the time of the
study in late June 1988. Based on a work site analysis,
workers were divided into potentially high and low exposure groups, and no statistically significant differences in
pulmonary function were found between the two groups,
after adjusting for smoking status.
A similar study was conducted by Hours et al. [53]: they
analysed 102 male workers employed by three French
urban incinerators during 1996, matched for age with 94
male workers from other industrial activities. The exposed
workers were distributed into 3 exposure categories based
on air sampling at the workplace: crane and equipment
operators, furnace workers, and maintenance and effluent-treatment workers. An excess of respiratory problems,
mainly daily cough, was more often found in the exposed
groups, and a significant relationship between exposure
and decreases in several pulmonary parameters was also
observed, after adjusting for tobacco consumption and
centre. The maintenance and effluent group, and the furnace group had elevated relative risks for skin symptoms.
In the same year, Takata et al. [54] conducted a cross-sectional study in Japan on 92 workers from a municipal
solid waste incinerator to investigate the health effects of
chronic exposure to dioxins. The concentrations of these
chemicals among the blood of the workers who had
engaged in maintenance of the furnace, electric dust collection, and the wet scrubber of the incinerator were
higher compared with those of residents in surrounding
areas, but there were no clinical signs or findings correlated to blood levels of dioxins.
In sum, there are some studies that suggest increased gastric cancer and respiratory problems among incinerators
workers. However, there are a great number of uncertainties, which make it difficult to derive conclusions.
Occupational studies on incinerator employees
Four studies conducted on incinerator employees were
reviewed (see additional file 3).
In 1997, Rapiti et al. [51] conducted a retrospective mortality study on 532 male workers employed at two municipal waste incinerators in Rome (Italy) between 1962 and
1992. Standardized mortality ratios (SMRs) were computed using regional population mortality rates. Mortality
from all causes resulted significantly lower than expected,
and all cancer mortality was comparable with that of the
general population. Mortality from lung cancer was lower
than expected, but an increased risk was found for stomach cancer: analysis by latency since first exposure indicated that this excess risk was confined to the category of
workers with more than 10 years since first exposure.
Epidemiological studies of health effects of other
waste management processes
Twelve epidemiologic studies on the potential adverse
health effects of other waste management practices are
reviewed and listed in additional file 4.
Waste collection
Ivens et al. [55] investigated the adverse health effects
among waste collectors in Denmark. In a questionnairebased survey among 2303 waste collectors and a comparison group of 1430 male municipal workers, information
on self-reported health status and working conditions was
collected and related to estimated bioaerosol exposure.
After adjusting for several confounders (average alcohol
consumption per day, smoking status, and the psychosocial exposure measures support/demand ), a dose-
Page 8 of 14
response relationship between level of exposure to fungal
spores and self-reported diarrhoea was indicated, meaning that the higher the weekly dose, the more reports of
gastrointestinal symptoms.
In contrast with these results, a study of 853 workers
employed by 27 municipal household waste collection
departments in Taiwan did not find an excess of gastrointestinal symptoms [56]. The workers answered a questionnaire and were classified into two occupational groups by
specific exposures based on the reported designation of
their specific task. The exposed group included those
working in the collection of mixed domestic waste, front
runner or loader, collection of separated waste and special
kinds of domestic waste (paper, glass, etc.), garden waste,
bulky waste for incineration, and the vehicle driver; the
control group included accountants, timekeepers, canteen
staff, personnel, and other office workers. No significant
differences were found in the prevalence of gastrointestinal symptoms, but results indicated that all respiratory
symptom prevalence, except dyspnoea, were significantly
higher in the exposed group, after adjusting for age, gender, education, smoking status, and duration of employment.
Composting facilities
In a German cross sectional study by Bünger et al. [57],
work related health complaints and diseases of 58 compost workers and 53 bio-waste collectors were investigated and compared with 40 control subjects. Compost
workers had significantly more symptoms and diseases of
the skin and the airways than the control subjects. No correction was performed for the confounding effect of
smoking, as there were no significant differences in the
smoking habits of the three groups.
A subsequent study in Germany by Herr et al. [58] examined the health effects on community residents of bio-aerosol, emitted by a composting plant. A total of 356
questionnaires from residents living at different distances
from the composting site, and from unexposed controls
were collected: self-reported prevalence of health complaints over past years, doctors' diagnoses, as was residential odor annoyance; microbiological pollution was
measured simultaneously in residential outdoor air.
Reports of airway irritation were associated with residency
in the highest bio-aerosol exposure category, 150-200 m
(versus residency >400-500 m) from the site, and periods
of residency more than five years.
Bünger et al. [59] conducted a prospective cohort study to
investigate, in 41 plants in Germany, the health risks of
compost workers due to long term exposure to organic
dust that specifically focused on respiratory disorders.
Employees, exposed and not exposed to organic dust,
were interviewed about respiratory symptoms and diseases in the last 12 months and had a spirometry after a 5year follow-up. Exposure assessment was conducted at 6
out of 41 composting plants and at the individual level.
Eyes, airways and skin symptoms were higher in compost
workers than in the control group. There was also a
steeper decline of Forced Vital Capacity among compost
workers compared to control subjects, also when smoking
was considered.
Materials recycling facilities
There are no epidemiological studies of populations living near materials recycling facilities; only studies on
employees are available.
In the already-quoted study by Rapiti et al. [51] on workers at two municipal plants for incinerating and garbage
recycling, increased risk was found for stomach cancer in
employees who had worked there for at least 10 years,
while lung cancer mortality risk was lower than expected.
In the study by Rix et al. [60], 5377 employees of five
paper recycling plants in Denmark between 1965 and
1990 were included in a historical cohort, and the
expected number of cancer cases was calculated from
national rates. The incidence of lung cancer was slightly
higher among men in production and moderately higher
in short term workers with less than 1 year of employment; there was significantly more pharyngeal cancer
among males, but this may have been influenced by confounders such as smoking and alcohol intake.
Sigsgaard et al. [61] conducted a cross-sectional study to
examine the effect of shift changes on lung function
among 99 recycling workers (resource recovery and paper
mill workers), and correlated these findings with measurements of total dust and endotoxins. Exposure to
organic dust caused a fall in FEV1 over the work shift, and
this was significantly associated with exposure to organic
dust; no significant association was found between endotoxin exposure and lung function decreases.
The same authors [62] also analysed skin and gastrointestinal symptoms among 40 garbage handlers, 8 composters and 20 paper sorters from all over Denmark, and
found that garbage handlers had an increased risk of skin
itching, and vomiting or diarrhoea.
In a nationwide study, Ivens et al. [63] reported findings
of self-reported gastrointestinal symptoms by selfreported type of plant. A questionnaire based survey
among Danish waste recycling workers at all composting,
biogas-producing, and sorting plants collected data on
occupational exposures (including questions on type of
plant, type of waste), present and past work environment,
Page 9 of 14
the psychosocial work environment, and health status.
Prevalence rate ratios adjusted for other possible types of
job and relevant confounders were estimated with a comparison group of non-exposed workers, and an association was found between sorting paper and diarrhoea,
between nausea and work at plastic sorting plants, and
non-significantly between diarrhoea and work at composting plants.
The health status of workers employed in the paper recycling industry was also studied by Zuskin et al. [64]. A
group of 101 male paper-recycling workers employed by
one paper processing plant in Croatia, and a group of 87
non-exposed workers employed in the food packing
industry was studied for the prevalence of chronic respiratory symptoms, and results indicated significantly higher
prevalence of all chronic respiratory symptoms were
found in paper workers compared with controls.
Gladding et al. [65] studied 159 workers from nine materials recovery facilities (MRFs) in the United Kingdom.
Total airborne dust, endotoxins, (1-3)-beta-D-glucan were
measured, and a questionnaire-survey was completed.
The results suggest that materials recovery facilities workers exposed to higher levels of endotoxins and (1-3)-betaD-glucan at their work sites experience various workrelated symptoms, and that the longer a worker is in the
MRF environment, the more likely he is to become
affected by various respiratory and gastrointestinal symptoms.
Choosing relative risk estimates for health
impact assessment of residence near landfills
and incinerators
The reviewed studies have been used to summarize the
evidence available, as indicated in table 1. When the overall degree of evidence was considered "inadequate" we
decided not to propose a quantitative evaluation of the
relative risk; when we arrived to a conclusion that "limited" evidence was available, relative risk estimates were
extracted for use in the health impact assessment process.
Table 2 summarizes the relevant and reliable figures for
health effects related to landfills and incinerators. For
each relative risk the distance from the source has been
reported as well as the overall level of confidence of the
effect estimates based on an arbitrary scale: very high,
high, moderate, low, very low.
Landfills
From the review presented above and following the work
already made by Russi et al. [11], it is clear that the studies
on cancer are not sufficient to draw conclusions regarding
health effects near landfills, both with toxic and non-toxic
wastes. The largest study conducted in England by Jarup et
al. [21] does not suggest an increase in the cancer types
that were investigated. Investigations of other chronic dis-
Table 1: Summary of the overall epidemiologic evidence on municipal solid waste disposal: landfills and incinerators.
HEALTH EFFECT
All cancer
Stomach cancer
Colorectal cancer
Liver cancer
Larynx cancer
Lung cancer
Soft tissue sarcoma
Kidney cancer
Bladder cancer
Non Hodgkin's lymphoma
Childhood cancer
Total birth defects
Neural tube defects
Orofacial birth defects
Genitourinary birth defects
Abdominal wall defects
Gastrointestinal birth defects§
Low birth weight
Respiratory diseases or symptoms
LEVEL OF EVIDENCE
LANDFILLS
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Inadequate
Limited
Limited
Inadequate
Limited*
Inadequate
Inadequate
Limited
Inadequate
INCINERATORS
Limited
Limited
Limited
Limited
Inadequate
Limited
Limited
Inadequate
Inadequate
Limited
Inadequate
Inadequate
Inadequate
Limited
Limited**
Inadequate
Inadequate
Inadequate
Inadequate
"Inadequate": available studies are of insufficient quality, consistency, or statistical power to decide the presence or absence of a causal association.
"Limited": a positive association has been observed between exposure and disease for which a causal interpretation is considered to be credible, but
chance, bias, or confounding could not be ruled out with reasonable confidence.
* Hypospadias and epispadias
** Renal dysplasia
§ The original estimates were given for "surgical corrections of gastroschisis and exomphalos"
Page 10 of 14
Table 2: Relative risk estimates for community exposure to landfills and incinerators
Health effect
Landfills
Congenital malformations [24]
All congenital malformations
Neural tube defects
Hypospadias and epispadias
Abdominal wall defects
Gastroschisis and exomphalos*
Low birth weight [24]
Very low birth weight
Incinerators
Congenital malformations [45]
Facial cleft
Renal dysplasia
Cancer [30]
All cancer
Stomach cancer
Colorectal cancer
Liver cancer
Lung cancer
Soft-tissue sarcoma
Non-Hodgkin's lymphoma
Distance from the source
Relative Risk (Confidence Interval)
Level of confidence**
Within 2 km
Within 2 km
Within 2 km
Within 2 km
Within 2 km
Within 2 km
Within 2 km
1.02 (99% CI = 1.01-1.03)
1.06 (99% CI = 1.01-1.12)
1.07 (99% CI = 1.04-1.11)
1.05 (99% CI = 0.94-1.16)
1.18 (99% CI = 1.03-1.34)
1.06 (99% CI = 1.052-1.062)
1.04 (99% CI = 1.03-1.06)
Moderate
Moderate
Moderate
Moderate
Moderate
High
High
Within 10 km
Within 10 km
1.30 (95% CI = 1.06-1.59)
1.55 (95% CI = 1.10-2.20)
Moderate
Moderate
Within 3 km
Within 3 km
Within 3 km
Within 3 km
Within 3 km
Within 3 km
Within 3 km
1.035 (95% CI = 1.03-1.04)
1.07 (95% CI = 1.02-1.13)
1.11 (95% CI = 1.07-1.15)
1.29 (95% CI = 1.10-1.51)
1.14 (95% CI = 1.11-1.17)
1.16 (95% CI = 0.96-1.41)
1.11 (95% CI = 1.04-1.19)
Moderate
Moderate
Moderate
High
Moderate
High
High
*The original estimates were given for "surgical corrections of..". **The following scale for the level of confidence has been adopted: very high, high,
moderate, low, very low.
eases are lacking, especially of respiratory diseases, yet
there is one indication of an increased risk of asthma in
adults [19], but with no replication of the findings. Overall, the evidence that living near landfills may be associated with health effects in adults is inadequate.
A slightly different picture appears for congenital malformations and low birth weight, where limited evidence
exists of an increased risk for infants born to mothers living near landfill sites. The relevant results come from the
European EUROHAZCON Study [23] and the national
investigation from Elliott et al. [24]. In the UK report, statistically significant higher risk were found for all congenital malformations, neural tube defects, abdominal wall
defects, surgical correction of gastroschisis and exomphalos, and low and very low birth weight for births to people
living within two km of the sites, both of hazardous and
non-hazardous waste. Although several alternative explanations, including ascertainment bias, and residual confounding cannot be excluded in the study, Elliott et al.
[24] provide quantitative effect estimates whose level of
confidence can be considered as moderate.
from socioeconomic status near the incinerators and a
concern of misdiagnosis among registrations and death
certificates for liver cancer. The histology of the liver cancer cases was reviewed, re-estimating the previously calculated excess risk (from 0.95 excess cases 10-5/year to
between 0.53 and 0.78 excess cases 10-5/year). We then
graded the confidence of the assessment for these tumours
as "moderate" with the exception of liver cancer (high)
since the misdiagnosis was reassessed and the extent of
residual confounding was lower. In the study by Elliott et
al. [30] no significant decline in risk with distance for
non-Hodgkin's lymphoma and soft tissue sarcoma was
found. However, the studies of Viel et al. [33] and Floret
et al. [34] conducted in France and the studies from
Comba et al. [39] and Zambon et al. [40] in Italy provide
some indications that an excess of these forms of cancers
may be related to emissions of dioxins from incinerators.
As a result, we provided effect estimates in table 2 also for
non-Hodgkin's lymphoma and soft tissue sarcoma as
derived from the conservative "first stage" analysis conducted by Elliott et al. [30]. We graded the level of confidence of these relative risk estimates as "high".
Incinerators
Quantitative estimates of excess risk of specific cancers in
populations living near solid waste incinerator plants
were provided by Elliott et al. [30]. We have reported in
table 2 the effect estimates for all cancers, stomach, colon,
liver, and lung cancer based on their "second stage" analysis. There was an indication of residual confounding
With regards to congenital malformations near incinerators, Cordier et al. [45] provided effect estimates for facial
cleft and renal dysplasia, as they were more frequent in the
"exposed" communities living within 10 km of the sites.
Other reproductive effects, such as an effect on twinning
rates or gender determination, have been described; however the results are inadequate.
Page 11 of 14
Conclusions
Additional file 2
We have conducted a systematic review of the literature
regarding the health effects of waste management. After
the extensive review, in many cases the overall evidence
was inadequate to establish a relationship between a specific waste process and health effects. However, at least for
some associations, a limited amount of evidence has been
found and a few studies were selected for a quantitative
evaluation of the health effects. These relative risks could
be used to assess health impact, considering that the level
of confidence in these effect estimates is at least moderate
for most of them.
Studies on incinerators. The data provided represent a brief description
of the studies on populations living near incinerators.
Click here for file
[http://www.biomedcentral.com/content/supplementary/1476069X-8-60-S2.XLS]
Additional file 3
Studies on occupational exposures among incinerators and landfills
workers. The data provided represent a brief description of the studies on
workers of waste management plants.
Click here for file
Most of the reviewed studies suffer from limitations
related to poor exposure assessment, aggregate level of
analysis, and lack of information on relevant confounders. It is clear that future research into the health risks of
waste management requires a more accurate characterization of individual exposure, improved knowledge of
chemical and toxicological data on specific compounds,
multi-site studies on large populations to increase statistical power, approaches based on individuals rather than
communities and better control of confounding factors.
Additional file 4
Studies on other waste management processes. The data provided represent a brief description of the studies on population living near plants
using waste management technologies different from landfills and incinerators.
Click here for file
List of abbreviations used
Acknowledgements
EU: European Union; INTARESE: Integrated Assessment
of Health Risks of Environmental Stressors in Europe;
NHL: non-Hodgkin's Lymphoma; OR: Odds ratio; TEQ:
Toxic Equivalent.
This study was funded by the INTARESE project. INTARESE is a 5-year
Integrated Project funded under the EU 6th Framework Programme - Priority 6.3 Global Change and Ecosystems. We thank Margaret Becker for a
linguistic revision the text. We are in debt to Martine Vrijheid for her comments on an earlier version of the manuscript.
Competing interests
The authors declare that they have no competing interests.
Authors' contributions
DP participated in the design of the study, conducted the
systematic review and drafted the manuscript. SM conducted the systematic review and contributed to draft the
manuscript. AIL participated in the systematic review and
contributed to draft the manuscript. CAP helped to conceive of the study and to write and revise the manuscript.
FF conceived and coordinated the study and helped to
write and revise the manuscript. All authors have read and
approved the final manuscript.
Additional material
Additional file 1
Studies on landfills. The data provided represent a brief description of the
studies on populations living near landfills.
Click here for file
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Page 14 of 14
Human Health and Chemical Mixtures:
An Overview
developed countries, Figure 1 shows a
chromatogram of urine extracted from
wet diapers of two infant boys 1 year of
age. The peaks in the chromatogram are
David 0. Carpenter, Kathleen F. Arcaro, Brian Bush,
identified single PCB congeners and the
dichlorodiphenyltrichloroethane metaboWilliam D. Niemi, Shaokun Pang, and Dilip D. Vakharia
lite dichlorodiphenyldichloroethylene.
Department of Environmental Health and Toxicology, School of Public
One infant (upper trace) was breast fed,
while the other (lower trace) was not. The
Health, University at Albany, Rensselaer, New York; Wadsworth Center,
analyses were conducted to determine
New York State Department of Health, Albany, New York
whether the breast-fed infant had more
Unlike laboratory animals, people are rarely exposed to a single hazardous chemical. However, PCBs and pesticides in his urine than the
most of the information documenting adverse human health effects from environmental and non-breast-fed infant, as breast milk
occupational contaminants has come from studies focused on exposure to single chemicals, and reflects the composition of lipophilic subthere is little information available on how two or more contaminants affect humans. Most stances present in a mother's body fat.
information on the effects of mixtures comes from animal systems and limited investigations of What is important is that these chroisolated human cells in culture, even though the study of mixtures in such systems has also been matograms show that both infants at 1
neglected. Two or more compounds may show additive, antagonistic, or synergistic interactions year of age already had significant evior may act on totally different systems and thus not interact. Furthermore, even a single chemical dence of exposure to a large number of
may have multiple effects and affect more than one organ system. Effects may vary with age, chemicals even though their dietary intake
and metabolites may have totally different actions from the parent compound. This paper will at this age was limited. These substances
review the variety of health effects in humans that may result from environmental contaminants are lipophilic and are retained in body fat.
and discuss how such contaminants may interact with each other. We will also present examples What is excreted in urine is only a small
on how different contaminants interact from toxicologic studies of polychlorinated biphenyls reflection of total exposure and body burperformed as part of our Albany, New York, Superfund Basic Research Program project. den. The question of importance for the
- Environ Health Perspect 106(Suppl 6):1263-1270 (1998). http.//ehpnet1.niehs.nih.gov/docs/
health of these boys is not simply what
1998/Suppl-6/1263-1270carpenter/abstract.html
chemical X does to their development and
Key words: metals, PCBs, estrogen disruptors, thyroid, cancer, birth defects, persistent
health, but rather what the impact is of all
organics, neurobehavioral effects
of these chemicals acting together.
Health Effects of Mixtur
Some of the major broad categories of
human diseases that are suspected to result
from exposure to environmental contaminants are cancer, birth defects, immune
system defects, reduced intelligence quotient (IQ), behavioral abnormalities,
decreased fertility, altered sex hormone
balance, altered metabolism, and specific
organ dysfunctions (2). Almost every
organ system may be affected by one or
more substances commonly found in our
environment. The diseases listed are
abstracted from many studies of both
humans and animals, and in most cases
these investigations were focused on a single contaminant. Some of these diseases,
when expressed in a given individual, are
difficult to ascribe with certainty to a particular exposure (3). This is true for
This paper is based on a presentation at the Conference on Current Issues on Chemical Mixtures held 11-13 cancer, birth defects, and many of the
August 1997 in Fort Collins, Colorado. Manuscript received at EHP 17 February 1998; accepted 24 June 1998.
Supported by National Institute of Environmental Health Sciences Superfund Basic Research Program grant endocrine disruptor and nervous system
actions. But others, such as the specific
P42 ES04913.
Address correspondence to D.O. Carpenter, School of Public Health, One University Place, Rensselaer, NY organ system dysfunctions seen with kid12144. Telephone: (518) 257-2025. Fax: (518) 525-2665. E-mail: [email protected]
ney disease following lead exposure (4), or
Abbreviations used: Ah receptor, aryl hydrocarbon receptor; CYP, cytochrome P450; DES, diethylstilbestrol;
E2, 17,1-estradiol; [,-gal, f-galactosidase; hER, human estrogen receptor; EC50, median effective concentration; the loss of particular neurons following
HxCB, hexachlorobiphenyl; IC50, concentration that inhibits 50%; IQ, intelligence quotient; LTP, long-term methylmercury exposure (5), are clearly
potentiation; PAHs, polycyclic aromatic hydrocarbons; PB, phenobarbital; PCB, polychlorinated biphenyl; PeCB, attributable to particular exposures. Many
pentachlorobiphenyl; PTU, propylthiouracil; 2,3,7,8-TCDD, 2,3,7,8 tetrachlorodibenzo-p-dioxin; TeCB, tetraof the effects of contaminants on humans
chlorobiphenyl; TEFs, toxic equivalent factors; Th, T helper; TrCB, trichlorobiphenyl.
Human exposure to environmental
contaminants is omnipresent. It does not
occur only in individuals who live next to
hazardous waste sites or just the disadvantaged and poor who live in inner cities or
third-world countries. Everyone carries a
burden of lead in his or her bones, mercury
in their hair, and dioxins and polychlorinated biphenyls (PCBs) in their body fat.
Environmental contamination is a global
issue, and contaminants in one country
often are transported to others via air,
water, foodstuffs, manufactured products,
and travelers. Environmental contaminants
may be natural substances such as metals or
radioactive materials, or they may be
manufactured products that are useful to
humans but still have toxic effects.
Contaminants may be manufactured or they
may be unintentional by-products of
human activity, as in the case of combustion
or incineration products or the generation
of chloroform as a result of chlorination of
drinking water. Many contaminants are
mixtures of related chemicals such as polycyclic aromatic hydrocarbons (PAHs), crude
and refined petroleum products, and polychlorinated aromatics; in some of these cases
even the individual components have not
been examined for toxicity.
As documentation of the widespread
degree of contamination, especially in
Environmental Health Perspectives * Vol 106, Supplement 6 * December 1998
1 263
CARPENTER ET AL.
or,
Figure 1. Gas spectrometry chromatograms of urine extracts from diapers from two infants at 1 year of age, showing presence of various single PCB congeners and pesticides. Analysis was made as described by Bush et al. (1).
The infant in the upper trace was breast fed, while that in the lower trace was not. Each peak is an identified PCB
congener or pesticide.
are subtle and difficult to quantify. This is
particularly true for alterations that occur
during development and are thus irreversible, as are many of the effects on the
nervous system and organs that are
hormonally regulated.
There are a number of factors that
complicate the toxicologic evaluation of
mixtures. Two or more compounds may
have additive effects as a result of acting at
the same site, altering the same process by
different mechanisms, or as a result of one
compound altering the metabolism of the
second in such a way as to generate a toxic
metabolite. They may also have antagonistic effects or may be synergistic in that the
two together give much greater response
than the sum of either alone. However,
they also may have absolutely no interactions, with each substance acting independently. In this case the net effect of the lack
of interaction is that a person experiences
the sum of the different organ toxicities of
the different contaminants.
There are several complications that
must be recognized when generalizing about
mixtures and coming from what we know
about how single compounds act. First, a
single compound may have multiple sites of
action and these may be mediated by totally
different mechanisms. Second, many substances, including metals, are changed to
metabolites or conjugates in the body, and
1264
these new products may also have biologic
activity that may or may not be similar to
the parent compound. Thus even a single
compound may become a functional mixture, as will be demonstrated for a single
PCB congener and its hydroxylated metabolites. Third, there may be different effects of
a single environmental contaminant at different ages. Lead is a clear example in that
levels of blood lead that appear to have little
effect on neurobehavioral function in an
adult can cause an irreversible decrement in
IQ and trigger altered behavior when
impacting the developing nervous system in
the prenatal or early postnatal period.
Environmental Diseases Resulting
from Genetic Damage
As shown in Figure 2, there are extensive
interactions among many of the various
organ systems, such that alteration of one
may influence the function of others.
Central to many of the influences on biologic systems are effects that occur at the
level of genes. Genes control almost everything, not just the eye color of our children. Cancer is a disease of genetic
disruption. Cancer results from mutations
in genes, some induced by a variety of
environmental factors and some inherited
from mutations in previous generations.
Cancer genes may either be such as to
promote generation of cancer (oncogenes)
perhaps more frequently, are mutations
that result in the loss of cancer suppressor
functions. Many different environmental
contaminants are carcinogenic, including
some metals and organics. Mutations can
cause birth defects, as normal development is under the control of genetics. But
there can be other kinds of effects of environmental contaminants mediated by
genetic dysfunction. During normal
development genes are activated or inactivated at different stages, usually under the
control of growth factors and hormones.
Environmental contaminants may interfere with this developmental process. For
example, many of the effects of diethylstilbestrol (DES), the estrogenic substance
given to many pregnant women some
years ago, were the result of altered
expression of genes regulating sexual functions (6). Genes regulate many aspects of
hormonal production, brain development
and function, immune system balances,
and organ physiology, as well as cancer
and birth defects.
Environmental Contaminants
and the Immune System
There are also many direct effects of
environmental contaminants on various
organ systems. The immune system, for
example, is suppressed by some substances
such as dioxin, coplanar PCBs, and PAHs
(7). But the immune system is affected very
differently by some metals, which promote
hypersensitivity, rashes, and autoimmunity
(8). An exciting developing area of investigation suggests that the dominance of
different populations of T helper (Th) lymphocytes is a major factor in an individual's
immune responsiveness. In individuals
with normal immunity, it appears that the
ThI lymphocytes predominate; these lymphocytes produce a particular profile of
cytokines. Individuals with hyperimmunity
(showing asthma, skin rashes, and autoimmune syndromes) have predominately
Brain
Hormones
- Genes 4
Reproductive
,< system
Immune
system
Figure 2. Diagrammatic interactions between genetic
information and various organ systems intimately
involved in the effects of environmental agents.
HUMAN HEALTH AND CHEMICAL MIXTURES
Th2 lymphocytes, which produce different
cytokines (9). It has recently been demonstrated that environmental contaminants
such as lead and mercury can alter the balance between the Thl and Th2 lymphocytes (10) and there is speculation that
contaminant exposure early in life may
cause permanent or at least prolonged
abnormalities in immune function.
The immune system is also intertwined
with the other hormonal systems and the
nervous system. Children exposed prenatally to DES show an altered immune
function (6). There is also extensive interaction between the nervous and immune
systems, even to the point of a common
use of messengers. Although neurotransmitters and cytokines have traditionally
been considered specific to only one system
or the other, we now know that both are
used at both sites (11).
Sex Steroids and the Nervous System
Estrogen has direct effects on neurons,
such that estrogen alters synthesis of some
transmitters (12), can alter neuronal structure (13,14), influences memory function
(15,16), can trigger taste aversions (17),
and can alter neuronal ionic currents (18).
Furthermore, estrogens protect neurons
against glutamate excitotoxicity (19).
Environmental Contaminants
and the Nenrous System
The nervous system is a frequent target of
toxic action. A number of organic and
inorganic compounds will cause abnormalities of peripheral sensory or motor
nerves, resulting either in loss of sensation,
abnormal sensation, or muscle weakness
(20). Since the studies of Needleman and
colleagues in 1979 (21), it has been
accepted that lead at remarkably low concentrations can cause a decrement in IQ
and also behavioral problems in children
exposed prenatally and in the early postnatal years. More recent studies have suggested that these actions are irreversible
(22). The exact mechanisms for these nervous system actions are not known.
Evidence from several laboratories suggests
that PCBs have similar effects; prenatal
exposure has resulted in decrements in
cognitive function and behavior (23,24)
that appear irreversible (25).
Polychioinated Biphenyls
as Chemical Mixue
Polychlorinated biphenyls are interesting
compounds for illustration of the multiple
effects of both individual chemicals and
mixtures; throughout the rest of this paper,
examples of results of PCB research from
our group in Albany, New York, will be
interposed. There are 209 possible PCBs
and 75 dioxins, depending on the number
and position of the chlorines on the base
biphenyl or dioxin rings. PCBs were made
as commercial mixtures with varying
degrees of chlorination. Although their
manufacture and use in the United States
ceased in 1977 when they were recognized
to be persistent both in the environment
and in the body, they continued until
recently to be manufactured and used in
many countries of the world. Although
they are persistent, they are altered by both
physical and biologic processes because
they are vulnerable to both anaerobic and
aerobic biodegradation (26-28) and
metabolism within the body (29). In most
cases these various forms of metabolism
alter the numbers and positions of the
chlorines and do not totally degrade the
PCB. Therefore, the number of different
chemical compounds that can affect
human and animal health is not limited to
the approximately 150 congeners that were
commercially produced.
Historically, PCBs have been considered
weak dioxins. Indeed, some PCB congeners
(those that can assume a coplanar configuration having chlorine atoms only in the meta
and para positions to the biphenyl bridge)
are weak activators of the aryl hydrocarbon
(Ah) receptor, which is known to mediate
many of the effects of 2,3,7,8 tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD),
the most toxic polychlorinated dioxin congener (30). On the basis of this assumption,
some have concluded that the PCBs that
cannot assume a coplanar configuration are
nontoxic (31), but this has now been
proven incorrect. Results from our laboratories and those of a number of other
researchers (32-42) clearly show that
many of the 209 possible PCB congeners
have discrete profiles of toxic actions and
each congener may indeed have multiple
actions at different sites.
ortho-substituted and lower chlorinated
PCBs are neurotoxic and at least three
different forms of direct neurotoxic actions
have been demonstrated. Shain et al. (32)
showed that congeners with two or more
ortho chlorines inhibit the enzyme tyrosine
hydroxylase, which is the rate-limiting
enzyme for the synthesis of the neurotransmitter dopamine. Kodavanti et al. (33) and
Carpenter et al. (34) showed that orthosubstituted, but not coplanar, congeners kill
cerebellar granule cells by a mechanism that
probably involves disruption of calcium
homeostasis. Finally, Niemi et al. (35)
demonstrated that both ortho and coplanar
congeners are capable of blocking the
process of long-term potentiation (LTP), an
electrophysiologic measure in the brain that
is thought to be correlated with cognitive
function (43).
In addition to the direct effects of
estrogens on neurons discussed above,
other endocrine systems are important to
nervous system function. One such system is the thyroid. The thyroid controls
the rate of metabolism and is essential to
organ development as well as daily function. Congenital hypothyroidism, even if
treated after birth, results in a syndrome
of minimal brain dysfunction (44). One
end of the structure of the thyroid hormone shows some steric features similar to
PCBs and dioxins and their hydroxylated
metabolites. These substances interfere
with normal thyroid function in a variety
of ways, as outlined in Figure 3. Several
Hypothalamus
T
TRH
Effects of PCBs
Pituitary
TSH
Increased
Thyroid
gland
Altered structure,
increased weight
T4
Decreased due to
T3
Binding
proteins
increased excretion
because of decreased
binding and increased
glucuronidation
Blocks binding,
(3-8x} than T4
Target Weak agonists
receptors and antagonists
Physical
growth and
metabolism
Reduced growth
and metabolism
Brain
development
Minimal brain
dysfunction syndrome
Organ
maturation
Abnormal
testicular
development
Figure 3. Diagrammatic indication of sites in which
polychlorinated biphenyls alter thyroid hormone function and influence development. Abbreviations: T3,
3,3',5'-triiodothyronine; T4, thyroxine; TRH, thyrotropinreleasing hormone; TSH, thyrotropin-stimulating hormone. Data from Porterfield (44), Byrne et al. (45), and
Visser et al. (46).
1 265
CARPENTER ET AL.
studies have shown that PCBs cause
hypothyroidism in animals (45), but there
are a number of possible tar1get sites that
have been identified and it iis likely that
different congeners act at diifferent sites
(46). Some PCB congeners do not alter
thyroid function (36) but thie pattern of
those that do and those that do not is not
as simple as positioning the chlorines on
the PCBs at sites comparable to those for
the iodines on thyroid hormonies.
Hypothyroidism results in reduced
cognitive function and we hav,e shown that
animals made hypothyroid (during postnatal development with the aigent propylthiouracil (PTU) have a reduceed LTP (47).
Figure 4 shows results from aln experiment
in which we tested whether tlhere was any
interaction between the reducction of LTP
induced directly by acute e)Kposure to a
PCB congener (2,4,4'-trichl orobiphenyl
[TrCB]) and that caused by c-hronic PTU
treatment. Clearly the effects are additive.
In this experiment the hyp othyroidism
was not secondary to PCB e:xposure, but
previous publications (36,4'5-47) document the fact that PCB exposi ure can cause
hypothyroidism. This observanion suggests
an important principle whenLconsidering
biologic responses to env ironmental
100a
80-
T
No 244
244
a-
E 60
W
40-
._E
:c
20-
.-
0-
||
T
contaminants: Even a single compound may
influence a particular outcome by additive
effects through totally different mechanisms.
In this case PCBs may reduce LTP by causing hypothyroidism, but in addition PCBs
may reduce LTP by a direct action such that
the net effect may be additive.
Endocrine disruption via interference
with sex steroid hormones is a topic of
intense interest, although it is not clear that
these actions are necessarily more significant biologically than those actions that
alter general metabolism via disruption of
thyroid function. Many different chemicals
show estrogenic, antiestrogenic, androgenic, or antiandrogenic activities (37,
38,48-50). Based to a great degree on the
human experiments with DES (6) as well
as on extensive information from wildlife
(51), the sex steroid endocrine disruptive
effects of xenobiotics have been suggested
as causes of the reported decline in sperm
count and general fertility (3), causes of
birth defects of the reproductive system,
contributors or causes of cancers of
endocrine systems, and the causative agents
for the perceived increase in alteration in
sexual preferences (52). Further studies
must be done, however, before it can be
assumed that these conclusions apply to
human populations.
Figure 5 shows the variety of ways in
which xenobiotics can alter sex steroid
function. Substances can mimic or antagonize endogenous hormones. They may alter
the rates of synthesis or metabolism of the
endogenous hormone or they may directly
or indirectly alter the expression of receptors for the steroid (53). These various
interactions may be complex in chemical
mixtures, with each individual compound
potentially having multiple actions and
different compounds in the mixture
potentially acting at different sites that
influence the same final outcome.
Figure 6 illustrates ways in which PCBs
influence estrogenic function. Those coplanar PCBs that, like 2,3,7,8-TCDD, activate the Ah receptor, cause the induction
of cytochrome P450s (CYP) of the CYPIA
and CYPIB gene families. These P450s
appear to catalyze the metabolism of many
PCB congeners and other aromatic moieties
such as endogenous hormones, including
estradiol. They are all estradiol hydroxylases
but insert the hydroxyl group at different
sites: CYPJAl at the 2 position and
CYPIB1 at the 4 position (54,55). Estradiol
can be oxidized at several positions, and the
products are reactive, rapidly metabolized
further, and excreted. Measurement of the
2- and 4-hydroxylated metabolites indicates the relative activity of the two forms
of P450. When metabolism of estradiol is
increased, functional levels fall and an
altered estrogenic function ensues. A number of the ortho-substituted PCBs, but not
the coplanars, produce a pattern of enzyme
induction similar to that elicited by phenobarbital (56). Although the precise biochemical mechanisms and protein factors
involved in this induction process are not
well characterized, elevated levels of
CYP2B, CYP2C, and CYP3A enzymes
result and have the same effect in increasing metabolism and excretion. These
enzymes are primarily expressed in the
liver, although there may be limited expression in other tissues. Elevated rates of
Induction of
-1 B
Ah receptor CCYP1A,-
Cytochromes P450
-20
Control
n=6
Substrates,
inhibitors
PTU
n=4
metabolism
Figure 4. Additive effects of hypothyroidism induced
by chronic treatment of developing animals with PTU,
as described by Niemi et al. (47), and acute exposure
of isolated hippocampal brain slices to 2,4,4'-TrCB on
LTP, recorded in hippocampal area CA1, as described
by Niemi et al. 135). Each bar represents the magnitude
and SEM of the increase in the population excitatory
postsynaptic potential (EPSP) induced by a tetanic activation of the synaptic input, which reflects LTP. Acute
incubation of the brain slice with 1 pM 2,4,4'-TrCB
resulted in about 50% reduction of LTP (right). Slices
obtained from animals exposed to PTU postnatally, as
described by Niemi et al. (47), also showed about 50%
reduction in LTP as compared to control slices (no 244).
When slices prepared from animals exposed to PTU
were acutely incubated in the PCB congener, there was
no LTP whatsoever (left, 244).
12Z66
Increased
Cvol"Xo;z^~~~~~~~~ndcto ofA;n
Exclusive oinding
at gene regulatory
elements
PCBs
/Induction of
CYP2B, -2C, -23A
Estrogenic or
antiestrogenic
PCB metabolites
IF
Estrogen receptor
*
Estradiol
Altered
estrogen
PB receptor
Altered estrogenic function
Figure 5. Ways in which different substances can
cause endocrine disruption.
Figure 6. Sites of action of polychlorinated biphenyls
in altering estrogenic function. PB, phenobarbital.
hepatic metabolism of estradiol are by 2,3,7,8-TCDD and by direct determiobserved in animals exposed to PCBs (57). nation of inhibition of cDNA-expressed
In contrast to CYP2B, CYP2C, and human CYPIBI by 3,3',4,4',5,5'-HxCB
CYP3A enzymes, CYP1A1 and CYPIBI in 3,3',4,4',5-PeCB. Thus these two
appear to be inducible in a number of congeners have opposing actions in this
extrahepatic tissues including breast, pathway-inducing mRNA for synthesis of
an enzyme that they then directly inhibit.
uterus, and pituitary (58,59).
Some of the metabolites produced,
Several other important conclusions
especially mono- and dihydroxy PCBs, come from this investigation. Previous
may have estrogenic or antiestrogenic studies of toxic equivalent factors (TEFs)
activities of their own (39). In addition to of PCB congeners relative to 2,3,7,8inducing P450s, however, some PCBs can TCDD have been conducted primarily
directly inhibit these enzymes. Although using rodents or rodent-derived cell lines.
some lightly chlorinated PCBs bind to the The values obtained from human cells
active site of the P450s and hydroxylation are quite different, and the highest TEF
of the compound occurs, some of the is about an order of magnitude less
more heavily chlorinated congeners bind than that derived from rodent studies.
but are difficult to hydroxylate, so they are However, the different human cell lines
very effective inhibitors. Finally, through also behave somewhat differently dependactivation of the Ah receptor, 2,3,7,8- ing on which P450s they express. Thus
TCDD and probably also coplanar PCBs the problem of extrapolation from animay have inhibitory effects on estrogen- mals to humans is complex. Also, the
regulated gene transcription by exclusive most potent PCB congener in stimulating
binding at gene regulatory elements found estradiol metabolism in this study is
in the 5' flanking regions of estrogen- 3,4,4',5- tetrachlorobiphenyl (TeCB), an
responsive genes. The ligand-bound Ah environmentally relevant congener that
receptor appears to disrupt the estrogen has not been previously identified as havreceptor-Sp 1 complex that is involved in ing high Ah receptor binding affinity nor
transcriptional activation of human assigned a TEF value (62).
Although most of the antiestrogenic
cathepsin D by interaction at an overlapping xenobiotic response element (60). actions of PCBs can be explained by effects
There may be similar negative regulation on estrogen metabolism, there is also the
of other estrogen-regulated genes by the clear possibility of action at estrogen recepAh receptor.
tors. Furthermore, the effects of PCB
Recent work in our laboratories using metabolites may be different from that of
cultured human cells (61) has demonstrated the parent compound. For example, Pang
how complex these interactions can be. and co-workers (40) found that 3,4,5Pang and co-workers measured CYPlAI TrCB was a potent inducer of CYPlAl
and CYPIBI mRNA in several human cell and CYPIBI mRNA and a promoter of
lines including MCF-7 cells, a human breast estrogen metabolism by both 2- and 4cancer line (40). They demonstrated that a hydroxylation. However, this compound is
number of coplanar congeners increased metabolized to a 4-biphenylol (Figure 7).
both CYPlAI and CYP1Bi mRNAs but Gierthy and colleagues (39,63) use the
ortho-substituted congeners did not. They MCF-7 cell focus assay to identify estrothen investigated estradiol metabolism by genic and antiestrogenic properties of
measuring levels of 2- and 4-methoxyestra- xenobiotics; results with this parent comdiol, produced through the action of cate- pound and its metabolite are shown in
chol-0-methyl transferase after estradiol is Figure 8. In this assay, 3,4',5-TrCB is
hydroxylated, by gas chromatography/mass antiestrogenic, probably as a result of
spectrometry in the media of exposed cells. induced metabolism of estradiol. However,
Although for some of the congeners there the 4-hydroxy metabolite, 3,4',5-TrCB-4was a good correlation between the mRNA OH, has no antiestrogenic activity but
levels and the degree of estrogen metabo- shows clear estrogenic activity. Thus a parlism, for others (especially 3,3',4,4',5,5'- ent compound and a metabolite may have
hexachlorobiphenyl [HxCB] and 3,3',4,4',5- diametrically opposite actions.
Figure 9 shows evidence that these
pentachlorobiphenyl [PeCB]), metabolism
was much less than otherwise expected. This processes occur in whole animals and again
difference reflected a direct inhibition of the emphasizes how the different organ systems
P450, which they showed by demonstrating are interconnected. Compound 3,3',4,4'that both of these congeners block the TeCB is a coplanar congener that is antielevation of estradiol metabolism induced estrogenic (37,38). This congener has no
effect on tyrosine hydroxylase activity (32).
However, 3,3',4,4'-TeCB is metabolized
to a hydroxylated compound that is estrogenic. When developing rats are exposed
to 2,2',4,4'-TeCB, an ortho-substituted
Cl
QQci
Cl
3,4',5-Trichlorobiphenyl
HI
3,4',5-Trichloro-4-biphenylol
Figure 7. Structure of 3,4',5-trichlorobiphenyl and its
metabolic product 3,4',5-trichloro-4-biphenylol.
A
n11
n M1
E2
100-
~
incn
E0
r
a) X
80-
i
60-
,ca.
CO
E
h".
3,4',5-TrCB-4-0
._
x1
E
4020-
20-L-$.
0
3,4',5-Trc
0/
1-12 10-11
10-T
10- 10-
10-7 10- 10-5
B
120 -
2a'c 100-
'0cm
0.
3,4',5-TrCB-4-OH
10-9M E2
cO 80E a)
oLU 60=-
F
LY 156758
E
._
3,4',5-TrCB
o E 20O
LL E
o-
v
-
/
0
10-12 10-11101 10
-10
10-7 10-6 10-5
Molarity
Figure 8. Estrogenicity and antiestrogenicity of 3,4',5trichlorobiphenyl and one of its metabolic products,
tested in the MCF-7 cell focus assay as described by
Gierthy et al. (39,63). (A) Dose-response relation for
foci formation induced by E2 and the biphenylol. Note
the lack of estrogenic activity of the parent compound.
3,4',5-trichloro-4-biphenylol is estrogenic at 5 pM and
is thus 50,000 times less potent than estradiol. (B)
Antiestrogenicity tested in the presence of 10-9 M E2.
The parent PCB is antiestrogenic, whereas the hydroxylated metabolic product shows no activity. Compared
to the specific estrogen receptor blocker LY 156758,
the parent PCB is approximately 100 times less potent.
1267
CARPENTER ET AL.
0.08 -
)
0.080
T
1_
_*
T
T
_
_
_
_
_
_
T
.E
_
_
_
_
0.02
_
_
_
_
0
0.02_
_
_
0
_
_
_
_
_
_
_
_
0.00 -
_
_
O O.1 1
3.4e3',4'-TeCB
0
1 10 20
2,4t2',4'-TeCB
Dose, mg/kg/day
Figure 9. Dopamine concentrations in rat frontal
cortex of animals exposed perinatally to various concentrations of 3,4,3',4'- or 2,4,2',4' tetrachlorobiphenyl,
as described by Seegal et al. (41). Note that although
the ortho-substituted 2,4,2',4'-TeCB resulted in a significant reduction in brain dopamine levels, the coplanar 3,4,3',4'-TeCB caused a significant increase in
dopamine levels. The former result is thought to be due
to direct inhibition of the rate-limiting enzyme for
dopamine synthesis (tyrosine hydroxylase), whereas
the latter effect is a result of the activity of the hydroxylated degradation product, which is estrogenic.
*p< 0.05, **p< 0.01, ***p 0.01.
1.47 1.28 1.10-
> 0.92= 0.73LU 0.550.37 0.18n nnU.U-L-
_14
_10
-5
0
5
10
14
Days
Figure 10. The normal variation of serum estrogen
levels during the ovulatory cycle. The area between the
upper and lower traces represents the range of E2 profiles in normally ovulating women. Data derived from
Diagnostic Products Corporation (68) and Thorneycroft
et al. (69).
congener that inhibits tyrosine hydroxylase
activity, there is a reduction in the level of
brain dopamine in the adults. However,
when animals are developmentally exposed
to 3,3',4,4'-TeCB, exactly the opposite
result is found; dopamine levels are
increased in adults, which may be due to
the estrogenic activity of the hydroxylated
metabolite of 3,3',4,4'-TeCB (41).
The recent report by Arnold et al.
(64) of synergistic actions of weak
environmental estrogens has focused
attention on the possibility that weakly
active and especially persistent substances
might have effects in combination that far
exceed those expected by simple addition
of effects. Although the results of this
study have been questioned by its authors
and others (42,65,66), there remains
some evidence from behavioral studies
that synergism does occur (67). Table 1
shows results from the study by Arnold et
al. (64) (columns 2 and 3) as contrasted
with those from Arcaro et al. (42)
(columns 4 and 5), using the MCF-7
focus assay and a competitive receptor
binding assay with purified recombinant
human estrogen receptor (hER). Arnold
et al. (64) used purified recombinant
receptor directly (column 3) or expressed
this receptor in yeast together with the 13galactosidase (>-gal) reporter gene (column 2) to test the estrogenlike activity of
hydroxylated PCBs and pesticides. Results
in column 2 and 3 show that both the
EC50 for ,B-gal activity and the IC50 for
binding the hER are significantly lower
for a combination of the two hydroxylated PCBs than for either compound
alone, indicating that in combination the
hydroxylated PCBs are significantly more
potent. Similar conclusions were drawn
for the pesticides dieldrin and endosulfan.
In the study by Arcaro et al. (42), in both
the MCF-7 cell focus assay and the hER
binding assay, hydroxylated PCBs alone
showed estrogenlike activity. However,
mixtures were not more potent, indicating
that no synergy occurred. Of the pesticides studied, only endosulfan was weakly
estrogenic and the combination with
dieldrin was not synergistic.
The most important question with
regard to weak environmental estrogens is
whether they interact with endogenous
estradiol. This is an important question not
only for women of reproductive age, who
have high and fluctuating estrogen levels,
but especially for children, postmenopausal
women, and men, whose estrogen levels are
low. Figure 10 illustrates the typical fluctuations of estradiol concentration during the
ovulatory cycle. Figure 11 shows results of
the estrogenic response in MCF-7 cells to
estradiol alone, tested at 10-12 to 10-8 M,
and with three different concentrations of
the estrogenic PCB metabolite 2,4,6-TrCB4'-OH. It is important to note that there is
no evidence of any synergistic effect of
estradiol and 2,4,6-TrCB-4'-OH on the
response of MCF-7 cells. We conclude that
different amounts of this estrogenic PCB
metabolite together with varying physiologic concentrations of estradiol do not
exhibit any synergism in the in vitro situation. However, this in vitro study does not
negate the synergism study in turtles (67),
although it does pose a challenge for building a stronger case that synergism between
environmentally relevant estrogenic
substances occurs in humans.
In summary, it is difficult to study
chemical mixtures because of the variety of
ways in which the components may interact. However, even single compounds can
have complex actions at multiple sites,
varying with stage of development, and
120° (D
100-
c0
80-
E
E'
LJ 60c4
0
Table 1. Estrogenic and synergistic effects of weak xenoestrogens and mixtures.
f-gal activity,
-0.E
hER binding,
development,
EC50 pMb
* E2 alone
5 x 10 6M 2,4,6-1
v5 x 10-7M 2,4,6-T
015 x 10OAM 2,4,6-T
0
hER binding2
Chemical
EC50 pMa
IC50 pMa
IC50 pMb
17,B-estradiol
0.0001
0.001
0.0003
0.0005
2,4,6-TCB-4'-OH
0.0070
0.055
0.22
0.079
2,3,4,5'-TCB-4'-OH
0.0180
0.12
0.72
0.015
2,4,6-TCB-4-OH and 2,3,4,5-TCB-4'-OH
0.0015
0.005
0.18
0.015
Dieldrin
>33
> 50
ND
ND
Endosulfan
>33
> 50
>10
ND
Dieldrin and endosulfan
0.092
0.324
>10
ND
Abbreviations; EC50, median effective concentration; IC50, concentration that inhibits 50%; ND, not detectable.
"Data from Arnold et al. (64). bData from Arcaro et al. (42).
1 268
40 -
MCF-7 focus
20 - I
0-
T//
-
0
I
-12
-11
-10
-9
-8
Log molarity of E2
Figure 11. The estrogenic effects of three concentrations of 2,4,6-trichloro-4'-biphenylol alone and
together with dose-response curves of E2 in the MCF-7
cell focus assay, as described by Arcaro et al. (42). The
data show no synergistic effect when the biphenylol at
various concentrations was added together with E2.
may become functional mixtures in the
body as a result of metabolism. In reality
people are exposed to mixtures and if we are
ever to understand the human diseases that
people develop from exposure to environmental contaminants, we must study and
understand the interactions that occur in
mixtures. At the same time, it will probably
not be possible to understand the complexity of mixtures without understanding the
mechanisms whereby individual contaminants and their metabolites act, recognizing
all of the problems associated with species
and organ specificities, age-dependent
actions, dose-response relationships, and
the enormous interdependence of the various organ systems that can lead to indirect
as well as direct effects.
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61. Pang S. Development and application of HPLC/MS and
GC/MS methodologies to determine the effect of polychlorinated aromatic compounds on estrogen metabolism in humanderived cell lines [PhD Dissertation]. The University at Albany
School of Public Health, Albany, NY, 1997.
62. Safe S. Polychlorinated biphenyls (PCBs): environmental
impact, biochemical and toxic responses, and implications for
risk assessment. Crit Rev Toxicol 24:87-149 (1994).
63. Gierthy JF, Lincoln DW, Roth KE, Bowser SS, Bennett JA,
Bradley L, Dickerman HW. Estrogen-stimulation of postconfluent cell accumulation and foci formation of human MCF-7
breast cancer cells. J Cell Biochem 45:177-187 (1991).
64. Arnold SF, Klotz DM, Collins BM, Vonier PM, Guillette LJ
Jr, McLachlan JA. Synergistic activation of estrogen receptor
with combinations of environmental chemicals. Science
272:1489-1491 (1996).
65. McLachlan JA. Synergistic effect of environmental estrogens:
report withdrawn. Science 277:462-463 (1997).
66. Ramamoorthy K, Wang F, Chen I-C, Norris JD, McDonnell
DP, Leonard LS, Gaido KW, Bocchinfuso WP, Korach KS,
Safe S. Estrogenic activity of a dieldrin/toxaphene mixture in
the mouse uterus, MCF-7 human breast cancer cells, and yeastbased estrogen receptor assays: no apparent synergism.
Endocrinology 138:1520-1527 (1997).
67. Bergeron JM, Crews D, McLachlan JA. PCBs as environmental
estrogens: turtle sex determination as a biomarker of environmental contamination. Environ Health Perspect 102:780-781
(1994).
68. Diagnostic Products Corporation. Publ PITK E2-F. Los
Angeles, CA:Diagnostic Products Corp., 1995
69. Thorneycroft IH, Mishell DR Jr, Stone SC, Kharma KM,
Nakamura RM. The relation of serum 17-hydroxyprogesterone
and estradiol-171B levels during the human menstrual cycle. Am
J Obstet Gynecol 111:947-951 (1971).
Vol 106, Supplement 6 * December 1998
Comments by Salford City Council on the Determination of an
Application for the Environmental Permit for Barton Renewable Energy
Plant
Ref No EPR/SP3234HY
Our ref: 95814
th
Date: 9 August 2012
This consultation refers to the draft Determination of Application decision
document and permit for the Barton Renewable Energy Plant, operated by
Peel Energy Limited and located in Trafford. The reference number is
EPR/SP3234HY. A list of the documents considered is provided in Appendix
1.
These comments are further to those submitted in 2011 and a meeting with
the Environment Agency in June 2012 to clarify points on the decision
document. A copy of our draft comments is attached in Appendix 2.
Background
Detailed comments were submitted to the Environment Agency on 9 March
2011 raising concerns about the impact of emissions from the plant. The
Environment Agency has carefully considered the issues raised, carrying out
their own assessment and other points by their in house modelling team, the
Air Quality Modelling and Assessment Unit (AQMAU). The conclusion and
recommendations can be found in the documents C704 and AQMAUC748/776-RP02.
Installation
Peel Energy Ltd have applied for a permit to operate a waste to energy plant
under the Environmental Permitting Regulations Section 1.1. The plant also
falls under the Waste Incineration Directive (WID). The plant burns waste
wood materials and the products of combustion, after being treated to reduce
emissions, are discharged through a 44.23 metre stack. These comments
relate to air quality and the impact on the local environment which is in an air
quality management area declared for exceedence of the annual nitrogen
dioxide (NO2). The main area of concern is the emission of nitrogen dioxide;
other pollutants are emitted and although not discussed here, it does not
mean that they are inconsequential to the environment or opinions express
elsewhere by members of the public or agencies.
Emissions Limits and Modelling Results
The WID sets a daily emission limit value of 200 mg/m3; Peel Energy
modelled the emissions at a lower limit of 125 mg/m3. The modelling predicts
where the maximum concentration from the stack will land and at selected
receptors in Salford and Trafford. The Environment Agency (Ref 1)
recommends that the annual mean values are corrected by 1.11 which has
been applied to the summary table below. As hourly mean values are
estimated over a shorter time period this correction does not need to be
applied. Table 1 shows predicted annual mean and hourly exceedence at the
point of maximum impact in Salford and the nearest receptor. In brackets is
the % exceedence of the air quality standard.
Table 1 Summary of annual mean results for air quality modelling,
Results µ/m3
Nitrogen dioxide
Annual Mean*
Hourly mean **
Point of maximum impact
2.19 (5%)
17.9(9%)
Nearest property Tindall
Street
0.77(2%)
3.55(2%)
* corrected by 1.11
** no correction
Air Quality Standard (AQS)
Annual mean 40 µ/m3
Hourly mean 200 µ/m3
Impact on Salford
The results in Table 1 show that the worst case exceedence is 5 % of the air
quality standards and 2 % at the nearest receptor in Salford. Salford has the
highest annual mean receptor and Trafford has the highest hourly mean value
of 9% at a property in Bent Street.
The installation is sited in an air quality management area for exceedences of
the annual mean NO2. The area is also significantly above the air quality
objective and EU limit value for nitrogen dioxide with monitoring results
indicating it experiences some of the highest concentrations in Greater
Manchester. In Greater Manchester there is a prevailing south-westerly wind
direction, resulting in the plume falling in the residential area in Barton and
adjacent wards. This will increase the roadside and background levels of NO2
in the immediate area. The cumulative effect of this and other developments
will result in increasing levels of background concentration.
Background Concentration
The Environment Agency submitted additional information dated 25 July
2012, received 31 July 2012 regarding background concentrations.
The data reviewed by the applicant and the Environment Agency considered
two nearby NO2 tubes, and real time monitoring sites Eccles, M60 and Bury
Radcliffe. The M60 site, being located by the M60 best represents the local
environment for the Barton plant. Results from Liverpool Road and the M60
have an annual mean above the recommended of 40 µ/m3 set by the national
and European legislation. Concentrations are 50% above the standard for
sites SA34 and the M60. The long term tends shown that NO2 emissions have
not decreased as expected from improvements in vehicle emissions.
Table 2 Summary of Background Concentrations,
Results µ/m3
Year
2009
2010
2011
3 Year Average
SA34 Liverpool Rd
(Diffusion Tube)
62.2
63.6
52.1
59
SA35 Trevor Rd (Diffusion Tube)
41.4
42.8
36.9
40
Salford Eccles
(Continuous Monitor)
39
42
33
38
Salford M60
(Continuous Monitor)
70
60
64
65
On the basis of these results it is difficult to determine a background
concentration that is representative of the locality, in an area that is
significantly over the AQS. The United Utility plant will add to the background
concentrating as indicated in Para 3.8 (Ref 2), adding further uncertainty to
the background concentrations and the calculation of the Process
Contribution (PC). This is a key parameter for informing the determination of
the permit. Its reliability due to the uncertainty in background data and
applying it to an area where the air quality standard is significantly exceeded
is not appropriate as it does not consider the overall environmental standards.
As a result the assessment by the Applicant and the Environment Agency
does not adequately consider the cumulative effect of the two plants, (Barton
Biomass and United Utilities combustion plant), which will result in the
background concentrations increasing.
With regard to the reference made, in the draft determination document, to
Defra Environmental Permitting Guidance – The IPPC Directive Part A(1)
Installations and Part A(1) Mobile Plant March 2009 version 21, we believe
that the wrong paragraph has been used for deciding whether the Permit
should be granted. Para 4.54 (from 2009 Defra Guidance) refers to the
contribution to the Environmental Quality Standard from Existing Plant. The
Barton Biomass Plant is of course a New Installation and therefore para 4.50
is more appropriate:4.502. For a new installation (or a substantial change to an existing
installation, where the effect of the change bears significantly on a Community
EQS), if environmental quality before the installation begins to operate meets
the requirements of a Community EQS, then this must remain so after the
1
2009 Version 2 guidance was updated in March 2010 to Version3 and then in March 2011 to
Version3 existing installation from 7/1/2013 and new installations from 7/1/2013. The Decision
Document uses the 2009 DEFRA guidance and to keep continuity we have used the same document.
2
Defra 2009 guidance
installation comes into operation. If the necessary ELVs cannot be met then
the permit must be refused. However, there may be ways to reduce emissions
from other sources in such a circumstance, thus rendering ELVs and other
permit conditions for the installation viably achievable. Where a new
installation would only make a minor contribution to a breach of a Community
EQS, it will normally be more desirable for regulators to work together to
control the other, main sources of pollution, thus ensuring the EQS is met.
Taking the above guidance into account this permit should NOT be granted
unless it can be proved that the contribution to the EQS from the plant is
“minor” and that other emissions sources are controlled through permit
conditions to ensure compliance with the EQS. The contribution from the new
United Utilities Plant has not been taken into account nor does the EA have
any control over the contribution made by the traffic using the motorway
network and link roads. The major contributor to the exceedences of the EQS
is traffic and consequently is outside the control of the EA and therefore by
granting this permit the EA will be condoning a creeping background level that
already exceeds the EQS.
The detail contained in Annex 5 Addendum to the draft decision does refer to
New Plant and quotes para 4.52 of the above guidance;
4.513. If a Community EQS is already being breached in a particular area,
then a permit should not be issued to any new installation that would cause
anything beyond a negligible increase in the exceedance. Again, however,
if it is clear that a combination of controls on the proposed installation and
measures to reduce emissions from other sources will achieve compliance
with the EQS, then the installation may be permitted.
There is disagreement with the EA’s interpretation of “negligible”. This,
“small” contribution is large enough to interfere with any headway that may be
gained by any proposed Low Emission Zone (discussed later). Again the EA
does not have the powers to reduce emissions from all other contributors,
other than the UU Plant and other permitted processes, to make this a
suitable location for the Biomass plant.
Furthermore as the European Commission has recently refused the UK’s
application to extent the NO2 compliance date, it is much more important to
prevent additional sources increasing background levels and delaying our
compliance.
United Utilities (UU)
The decision documents (p59) indicates a 1-2 % increase in the annual mean,
concluding that the contribution will be negligible due to the conservative
nature of the modelling. However overall this will add to the total NO2
concentration in the area making it harder to achieve the standard. The
3
Defra 2009 guidance
conservative approach is also weakened by the proximity of the plants and
the overlap of the plumes, (which are not shown).
The accumulative affect of Barton and UU range from 5% to 7% in the worst
case condition and this is not adequately considered in the decision document
and neither is its effect on background levels included in the PC calculation.
This is relevant, given the issues raised in the previous section and the
proximity of the background concentration to the air quality objective. The
inclusion of this data will further reduce the headroom available.
Regarding the short term hourly objective there are recorded of exceedences
at the M60 which is not considered by the applicant even though it is operated
to same standard as Glazebury and Eccles. Also where tube concentration
exceeds 60 µg/m3 as at SA34, hourly exceedences may occur. Due to the
proximity of the stacks, it cannot be guaranteed that the two plumes, under
the right weather conditions will contribute or cause exceedences of the
hourly standard.
A more rigours approach is therefore required to access the impact of the two
plants.
Meeting European Commission Limit Values
The increased levels of NO2 will cause the size of the air quality management
area to increase and also delay compliance with the objective. The costs of
these delays by non compliance may result in fines from the EU being
cascaded to local authorities by central government .
Since the publication of the decision document, the European Commission
has refused the UK government an extension of the deadline to 2015 for the
achievement of annual mean NO2. This places greater emphasis and urgency
on the UK government to develop plans to attain the target values. The EU
also states that to due to public health impacts postponements should be as
short as possible, suggesting that where possible we should not be delaying
attainment where the AQS is breeched.
The local authorities and other regulating bodies such as the Environment
Agency and Highways Agency will have to work with the UK government to
develop effective plans to reduce NO2 levels and attain the European and
national standards as soon as possible.
The contribution from both plants, Barton Biomass at 5% and United Utilities
at 1-2% and therefore may be as high as 7%, and are a significant source of
new emissions.
Significance of the Plant’s Emission on Complying with European Limit Values
DEFRA prepared, as part of the submission for the EU extension, the
improvement to be gained if a Low Emission Zone (LEZ) were to be
introduced in Greater Manchester and other affected areas across the UK. In
this scenario all HGVs / buses would be required to meet Euro IV emissions
standards for NOx and PM10 in 2015 and would apply to most roads in a
Local Authority but not strategic roads like the M60. Comparison of the
scenarios prepared in the DEFRA report for this region indicates a gain of 1.6
µg/m3 in 2015; therefore the additional impact of 0.69 µg/m3 (0.76 µg/m3 if
factored by 1.11) is a significant portion of the projected gain being taken
away and subsequently weakening national and local air quality plans. The
United Utilities emissions will place additional constraints on meeting the
target.
Health Impacts
The areas affected by the plume are some of the poorest areas in Salford,
with a high deprivation index. Due to the high background concentrations the
impact of the emissions will have a greater affect in these areas than an
emission of similar size in an area where air quality is below the air quality
standard.
The costs assessment concludes that the additional reduction achieved in
NO2 in the AQMA is negligible. The assessment fails to include the cost of
NO2 emissions to the environment, action plans to reduce emissions and any
fines from the EU. DEFRAi provide a calculator to estimate the monetised
impact of emissions.
The annualised cost of abating NOx is presented in Table 3 for two different
abatement techniques. Selective Catalytic Reduction (SCR) reduces the
maximum contribution by 44% but at a higher operating cost however the
cheaper option is preferred. The assessment does not consider heath costs,
environmental damage or the costs of any fines by not meeting the EU
standard and should be included to show all the cost in the cost benefit
analysis. If included this will reduce the margins between the two techniques
thereby justifying the use of SCR.
Table 3 Annualised Cost of Abatement System
Technique
SCR
SNCR
Annualised Maximum
Contribution
Cost4
µg/m3
£748,121
1.1
£192,575
1.97
Difference
µg/m3
%
Difference
0.87
44
Furthermore the 0.87 µg/m3 reduction achieved with SCR is significant when
compared with the gains predicted if a Low Emission Zone (LEZ) is
implemented. The cost, based on per NOx reduced, of implementing an LEZ
is likely to significantly outweigh the additional cost incurred by using SCR.
Overall the additional cost of implementing SCR is not excessive compared to
fully costed assessment incorporating health impacts and action plan
4
Source: Decision Document p 75
measures to mitigate these reductions. It is therefore recommended that, if
the EA are minded to permit this process, SCR is included as the preferred
abatement technique as part of the additional measures to achieve Best
Available Technique (BAT) and to meet the statement set out in p 35 of the
Decision Document which states
“Additional measure will also be included in the process design to control the
emissions to a level significantly below that required by the WID.”
Particulate Analysis
It is recommended that particulate analysis is undertaken on regular intervals
and when these is a significant change in the feed stock. The following
species should be reported PM 10, PM2.5, PM1.0 and PM0.1. (See 5.24 p 36
Ref 4)
Arsenic
Arsenic emissions data reported have a wide range of values, (3.91 µg/m3 to
40.47 µg/m3, at Wilton) suggesting a large uncertainty in the results due to
sampling technique or the impact of different raw materials. Further
clarification on the source of the variance is needed. (See 5.2.25 p 36)
Scientific Evidence
The Audit commission report that over 50,000 people a year may die
prematurely from poor air quality. This should be included in the review of
evidence available. (See 5.3 p 39-42 Ref 4)
Port Salford / Salford Reds Stadium
The air quality assessment reportii prepared by AQC on behalf of Peel
modelled air quality in the area for 2010 by predicting changes in emissions
from a base year of 2005. The factors used for this have now been shown to
be over optimistic and it is widely acknowledged by DEFRA that long term
trends have not fallen as predicted.
Nonetheless the report indicated that the worst affected area would see an
increase of up to 1.2 µg/m3. The concentration predicted for SA34 in 2010 is
52 µg/m3 versus a measure concentration of 64 µg/m3. The model has
therefore overestimated the reductions in emissions.
The local authority are responsible for updating and incorporating the AQMA
which is undertaken by running a county wide air quality model at periodic
intervals at a time suitable to all 10 Greater Manchester districts.
This statement is therefore misleading and should be amended.
(See 5.6.2 p60 ref 4)
Previous Comments
There are a number of concerns regarding this application, raised previously,
including among other concerns:
•
•
•
Modelling: surface roughness,
The effect of Barton Bridge.
The location of a operating plant in an area where air quality standards
are significantly exceeded with some of the highest levels in greater
Manchester
The work done by the AQMAU has shown that surface roughness does not
significantly affect the outcome and we are in agreement with this.
The Barton Bridge is considered to have no impact on the plume from the
stack but no supporting evidence is provided for this statement. Further
justification is required to demonstrate that it does not interfere or affect the
plume trajectory.
Conclusions
The location of the plant and the impact of the emissions remain a serious
concern in terms of health impacts and compliance with the air quality
standards.
The Environment Agency’s assessments provide an accurate and
comprehensive appraisal of our comments and the applicant’s proposal. A
number of points have been clarified and adequately dealt with in reference 2
and 3, however some remain justified and the broad issues raised by the
operation of the plant therefore still exist. These are:
•
•
•
•
•
•
that the plant worsens air quality in the area that already has poor air
quality and is likely to continue to do so for some time,
it prevents the fulfilment of Salford’s air quality action plan
it contributes to the exceedences of the EU air quality limit values for
nitrogen dioxide and delays its attainment
it further delays the attainment of the nationally set air quality standard
for nitrogen dioxide to protect human by the UK government,
it’s likely to contribute to the number of hourly exceedences of the air
quality objective nitrogen dioxide
a number of technical issues and opinions that are different to
Environment Agency’s remain which should be addressed.
Recommendations
It is recommended that the permit is not granted. The EA should give further
consideration to DEFRA Environmental Permitting Guidance for Part A(1)
Installations, in particular para 4.46 – 4.55. Para 4.495 promotes co-operation
between regulators and the aim of improving areas of poor environmental
5
Defra 2009 guidance
quality. By allowing this permit the EA are not following this aim. Para 4.50
refers to new installations and that the new installation should only be
permitted if it makes a minor contribution and that regulators can work
together to control the other main sources, thus ensuring that the EQS is met.
There is a disagreement that this installation is a minor contributor and the
fact that major contributor, the motorway network, is out of the control of both
the EA and the Local Authority, where the LEZ, if implemented, will have
limited effect on vehicle emissions from the M60.
It is also recommended that the EA give greater weight to the uncertainty in
the background concentrations of NO2 versus the actual air quality with no
significant decline and the consequential heath effects.
Should the EA decide not to follow our recommendation and grant the permit
precedence should be given to technologies e.g. that will give the largest
reduction in NO2 emissions to air as area where there is the greatest
uncertainly and health impacts. It will also help reduce a creeping
background. Selective Catalytic Reduction (SCR) gives lower NOx emissions
and this or better techniques should be specified in the permit. This is in
accordance with para 4.486 of the above guidance.
It is also recommended that the Operator should supply annually mass
emissions, in addition to others, for NO2, PM10, PM2.5 and PM.1
Periodic particulate monitoring for the fractions PM10, PM2.5, PM1 and
PM0.1 should be undertaken at regular intervals or when there is a change of
feedstock.
For comments please contact
Lynda Stefek
Gerard Steadman
i
http://uk-air.defra.gov.uk/library/reports?report_id=639
Port Salford Environmental Statement 2nd Supplement – Vol. 2 and 3: date 27.07.06
ii
6
[email protected]
[email protected]
Defra 2009 guidance
Appendix 1 Document List
No
1
Name
S1100-0011-0009SMO AQA
BREP rev4.pdf
2
C704 Barton REP RP01.pdf
(17.01.2011)
C748-776 Barton REP RP02.pdf
(12.5.2011/28.6.2011)
3
4
Barton-Permit_DD-SP3234HYDRAFT_decision-May_12.pdf
5
Review of Background Air
Monitoring Data
(25.07.12)
Details
Peel Energy Environmental
Assessment Document Rev
4
Response to NPS audit
request
AQMAU comments on
Applicant’s response to first
audit
Determination of an
Application for an
Environmental Permit under
the Environmental Permitting
(England & Wales)
Regulations 2010
Appendix 2
Comments on the Determination of an Application for the Environmental Permit Ref
No EPR/SP3234HY (used at meeting 25/6/12 with Environment Agency)
Our ref: 95814
This consultation refers to the determination of application on permit reference no
EPR/SP3234HY
Detailed comments were submitted to the Environment Agency dated 9 March
2011 raising concerns about the impact of emissions from the plant.
Since these comments have been submitted additional work has been undertaken
by either the Environment Agency and or the applicant, Peel Energy Ltd and
information exchanged. Theses reports, listed below are available on the public
register, but were not available at the time of writing. Copies of the information
have been requested
11 May 2011
(Ref: S1100-0011-0021AMW)
18 January 2012
(Ref: S1100-0420-0051RSS)
2 March 2012
Email from Fichtner
AQMAU report (Ref: C704, dated
21/03/11)
AQMAU report (Ref:AQMAUC748/776-RP02), dated 23/08/11)
To clarify abnormal emissions
There are a number of concerns regarding this application, raised previously,
including:
•
•
•
Modelling: surface roughness,
The effect of Barton Bridge.
The location of a operating plant in an area where air quality standards are
significantly exceeded with some of the highest levels in greater
Manchester
None of the above have been adequately consider in this report.
The operation of the plant will increase the air quality management area and delaying
compliance with the EU and national standards. The costs of these delays by non
compliance may result in fines from the EU being cascaded to local authorities.
Further evidence and information is required to substantiate the statement made in
section 5.1.2. that emissions at the permitted limits would ensure a high level of
human protection, for the reason outline below:
Section/page Comment
5.2.1./29
The site is referred to as flat when there is a bridge and other large
buildings at the same height that may impact on the modelling.
Turbulence from the bridge will affect the trajectory of the plume and
areas affected, so the effects of local dispersion have not been taken
into account. Insufficient information given to justify this statement.
Further evidence e.g. Papers on previous studies or wind tunnels
studies are suggested.
5.2.2./30
5.2.2. /
p34/5
5.2.4/36
5.2.5
5.3/39-42
No information is provided on surface roughness and the effect of it.
The original study selected a low value for a rural area. i.e. choice not
justified given complexity of territory.
The Agency’s decision to adopt a conservative approach for the
conversion of NOx toNO2 of 70% is supported.
The applicant reports a contribution of 1.7% at worst case residential
receptor (location not given) but the maximum impact predicts an
increase of 1.97 µg/m3 µin the annual mean, a 5% change. ( Table
12.21)1
There is a considerable range of values possible.
The report refers to Para 4.48 and 4.54 of the Defra Guidance2Part
A(1) regarding the impact of local IPPC process. No information on
local process and there respective contributions is provided. National
information on PM10 emissions is provided on page 46. Trafford
Park and surrounding area has a large number of industrial processes
emitting NOx and particulates and therefore greater weight given to
IPPC processes in this area.
Recommended that particulate analysis size distribution is periodically
undertaken and when there is a substantial change in raw material.
Arsenic emissions data reported have a wide range of values, (3.91
µg/m3 to 40.47 µg/m3, at Wilton) suggesting a large uncertainty in the
results due to sampling technique or the impact of different raw
materials.
The studies mainly examine the effect of cancer and do not consider
other effects. Evidence from the Audit Commission states that
...”Poor air quality reduces the life expectancy of everyone in the UK by an
average of seven to eight months and up to 50,000 people a year may die
prematurely because of it”. This evidenced should form part of the
5.6.1 /
5.6.2 /59
5.6.2. /60
assessment.
Recent trends at Eccles have seen an increase in NO2 see information
below. Averaging results reduces the maximum impact which is 42
µg/m3 in 2010. The increasing emissions due to Part A process will
delay the attainment of the standards.
Cumulative impact of Carrington and United Utilities (UU) has been
considered by an audit of the air quality reports. The impact of the
Carrington plant will be less significant that Unitised Utilities which is
adjacent to the installations The conclusion reached is that the NO2
concentration will increase by 1-2%. Is this from the UU site or the
impact of both sites together? 1-2% exceeds the significance long term
criteria and therefore the assessments should be based on modelling
not auditing.
The Environment statement for Port Salford concludes that these
developments would not lead to amendments of the AQMA from the
report. This is not our understanding of the assessments neither is it
our understanding of how acknowledged increases in emissions do not
lead to an increased in local concentrations and therefore to an
1
BREP Environmental Statement Volume 1
2
Part A(1) installations
5.6.2. /60
6.2.2./74
6.2.2/75
increased / change in the AQMA. This statement needs to be
reassessed. Misleading?
The impact of the Biomass plant is taken at the maximum impact of
Port Salford. The impact should report the maximum impact of the
biomass emissions on Liverpool road which may be different.
There are concerns about the ability of the applicant to meet the tough
NOx emission standards of 125 mg/Nm3 (WID Limit is 200 mg/Nm3).
The letters of support by boiler manufactures all suggest a high level
of complexity and bespoke design to achieve these standards. Process
controls will need to be precise and responsive to meet the standards.
There remains high level of uncertainty that the limits will be met.
Will the techniques suggested affect the plume buoyancy and
trajectory and has the modelling adequately consider these different
systems particular during abnormal conditions and start-up/shutdown
The costs assessment concludes that the additional reduction achieved
in NO2 in the AQMA is negligible. The assessment fails to include
the cost of NO2 emissions to the environment. (3See links below). The
costing should include all costs especially those calculated from the
Inter Departmental Group on the Costs and Benefits of Air Quality Damage Cost Calculator. Not withstanding the environmental damage
the full impact to health has not been included in an area which levels
exceed the annual limit by 50% and hourly exceedences are highly
likely. The cost model should take account of the modelling
uncertainties and the range of concentrations possible e.g. maximum
ground level concentration in NO2 from the combined impact of UU
and the installation for a full understanding.
Table 1 Nitrogen Dioxide Results Automatic Stations (µg/m3)
Year
2008
2009
2010
2011
Glazebury
AURN
17.3
16
19.4
18.3
Salford Eccles
AURN
36
39
42
33
Salford M60
Cal Club
68
70
60
64
Trafford
Cal Club
32
34
33
26
Trafford A56
Cal Club
46
44
46
41
Rolling average ( 3 years)
Eccles
3
2006/8
35
2007/9
36
2008/10
2009/11
39
http://uk-air.defra.gov.uk/reports/cat19/1102150857_110211_igcb-damage-cost-calculator.xls
http://uk-air.defra.gov.uk/library/reports?report_id=639
38
Policy Analysis
pubs.acs.org/est
Public Health Impacts of Combustion Emissions in the United
Kingdom
Steve H. L. Yim and Steven R. H. Barrett*
Department of Aeronautics and Astronautics, Massachusetts Institute of Technology, Cambridge, Massachusetts, United States
S Supporting Information
*
ABSTRACT: Combustion emissions are a major contributor to degradation
of air quality and pose a risk to human health. We evaluate and apply a
multiscale air quality modeling system to assess the impact of combustion
emissions on UK air quality. Epidemiological evidence is used to quantitatively
relate PM2.5 exposure to risk of early death. We find that UK combustion
emissions cause ∼13,000 premature deaths in the UK per year, while an
additional ∼6000 deaths in the UK are caused by non-UK European Union
(EU) combustion emissions. The leading domestic contributor is transport,
which causes ∼7500 early deaths per year, while power generation and
industrial emissions result in ∼2500 and ∼830 early deaths per year,
respectively. We estimate the uncertainty in premature mortality calculations at
−80% to +50%, where results have been corrected by a low modeling bias of
28%. The total monetized life loss in the UK is estimated at £6−62bn/year or
0.4−3.5% of gross domestic product. In Greater London, where PM concentrations are highest and are currently in exceedance
of EU standards, we estimate that non-UK EU emissions account for 30% of the ∼3200 air quality-related deaths per year. In the
context of the European Commission having launched infringement proceedings against the UK Government over exceedances
of EU PM air quality standards in London, these results indicate that further policy measures should be coordinated at an EUlevel because of the strength of the transboundary component of PM pollution.
1. INTRODUCTION
Poor air quality adversely impacts human health.1,2 In
particular, long-term exposure to fine particulate matter results
in an increased risk of premature mortality,1,3−5 with the
likelihood of a causal link estimated at 90%.6,7 Although other
anthropogenic air pollutants are known to cause adverse health
impacts, long-term exposure to PM2.5 (particulate matter with
an aerodynamic diameter of less than 2.5 μm) is understood to
be the air pollution exposure metric that is most consistently
and independently associated with early death, and which
accounts for the majority of the health costs of air pollution.1,4
In 2009, the Committee on the Medical Effects of Air
Pollutants (COMEAP) estimated that there were ∼29,000 early
deaths in the UK in 2008 due to anthropogenic PM air
pollution.4 This corresponds to ∼340,000 life-years lost per
year.4 The COMEAP approach was based on a combination of
modeling and measurements of PM concentrations with a
scheme designed to achieve “mass closure” relative to measured
concentrations. Stedman et al.8 developed the method applied
by COMEAP to map PM concentrations across the UK at
background and roadside locations by summing modeled and
empirical components.
In addition to the aforementioned health impacts, the UK is
currently not in compliance with the legally binding EU PM air
quality standard on account of exceedances in London.9
Because of this, the European Commission has recently
launched infringement proceedings against the UK Govern© XXXX American Chemical Society
ment for this continuing breach, with the potential for
unlimited fines subject to ruling by the European Court of
Justice.9
Results of Whyatt et al.10 indicate that emissions control of
primary PM alone would not be sufficient to meet European
Union (EU) limit values for PM concentrations. Andersson et
al.11 estimated the contribution of different European regions to
population PM exposure and premature mortality. In the UK,
the costs and benefits of the past and potential mitigation
policies for electricity generation and road transport were
estimated using a combination of dispersion modeling and
empirical components for secondary PM.12 While these studies
have added to understanding of PM concentrations in the UK,
the attribution of air quality-related premature mortalities to
different sectorsboth within the UK and from the rest of the
EUhas not previously been quantified.
The primary control the UK has on its PM concentrations is
by influencing domestic (i.e., within the UK) combustion
emissions, although a significant fraction of impacts may be
transboundary.11 The predominant sources of anthropogenic
PM pollution are combustion emissions of primary PM and
precursors of secondary PM. Here we estimate the number of
Received: November 12, 2011
Revised: March 12, 2012
Accepted: March 21, 2012
A
dx.doi.org/10.1021/es2040416 | Environ. Sci. Technol. XXXX, XXX, XXX−XXX
Environmental Science & Technology
Policy Analysis
Simulated baseline concentrations are validated against UK
National Air Quality Archive time series data from 79 O3, 79
NO2, 61 PM10, and 4 PM2.5 measurement stations. See Section
3 of the SI for further information on the air quality simulation
and validation.
2.3. Health Impacts. The relationship between long-term
exposure to particulate matter and health has been quantified in
epidemiological studies.21−26 These studies have consistently
found that long-term exposure to particulate matter less than
2.5 μm in diameter (i.e., PM2.5) is associated with increased risk
of premature mortality. Assessing long-term PM2.5 exposure is
thought to capture ∼80% of monetized health impacts of air
pollution.27 We therefore focus on long-term PM2.5 exposure
and associated increased risk of premature mortality.
Concentration−response functions (CRFs) relate changes in
PM2.5 exposure to changes in premature mortality risk. A U.S.
EPA expert elicitation study reported a 1% (with a range of
0.4−1.8%) decrease in annual all-cause deaths per μg/m3
decrease in annual average PM2.5 exposure.1 Results are similar
to an EU expert elicitation study.28 We apply the EPA CRF to
estimate early deaths in UK due to long-term exposure of
sector-attributable PM2.5 for adults over 30 years of age. The
UK all-cause baseline death incidence rate is based on WHO
Health Statistics and Health Information Systems (2004).
Population density data are derived from the Gridded
Population of the World (GPWv3) at a 2.5′ resolution.29
domestic early deaths per year attributable to UK combustion
emissions from sectors including power generation, commercial, institutional, residential and agricultural sources, industry,
and transport. We also quantify the contribution of non-UK EU
emissions to air quality-related deaths in the UK, and vice versa.
The purpose of this is to inform UK and EU air quality and
emissions policy development.
2. MATERIALS AND METHODS
Our overall approach is to derive a temporally, spatially, and
chemically resolved emissions inventory for the UK (at high
resolution) and the EU (at low resolution) suitable for use in a
state-of-the-science atmospheric chemistry-transport model.
We evaluate meteorological and baseline air quality simulations
to quantify the extent to which the modeling approach
reproduces observed meteorological fields, total PM, and
other species concentrations due to all emissions. Scenarios
are modeled in which each sector’s combustion emissions are
removed in-turn. We attribute the difference between the
resultant PM concentrations for each sector simulation and the
simulation of total PM to the respective sector. Nonlinearities
are assessed for UK emissions by modeling a case where all UK
combustion emissions are removed. The impact of UK
combustion emissions on EU air quality and vice versa are
also simulated. Premature mortality impacts of each sector are
estimated by overlaying sector-attributable PM concentrations
onto population density, and multiplying the resultant exposure
by a concentration−response function.
2.1. Emissions. Emissions in the UK are derived from the
2007 National Atmospheric Emissions Inventory (NAEI),13
which has a horizontal resolution of 1 km ×1 km. In the rest of
Europe, the 2007 European Monitoring and Evaluation
Programme (EMEP)14 inventory is applied for area sources.
EMEP has a horizontal resolution of 50 km ×50 km. Point
source emissions outside the UK are from the European
Pollutant Release and Transfer Register (E-PRTR).15 We
calculate plume rise in-line for each point source according to
Briggs et al.16
Emissions are divided into United Nations Economic
Commission for Europe (UNECE) source categories (“sectors”
in this paper): (a) power generation; (b) commercial,
institutional, residential, and agricultural sources; (c) industry;
(d) road transport; and (e) other transport.
Uncertainties in the emissions inventories, and temporal and
vertical emissions profiles are described in Section 1.4 of the
Supporting Information (SI), along with assumed chemical
speciation profiles for VOCs, NOx, SOx, and primary PM
emissions.
2.2. Meteorological and Air Quality Modeling. The
Weather Research and Forecasting Model (WRF)17 is used to
derive meteorological fields, driven by six-hourly ECMWF
reanalysis for the year 2005.18 European meteorology is
simulated at a resolution of 40.5 km, with a two-way nest to
a 13.5 km domain encompassing the UK. Meteorology is
validated with reference to 106 wind stations and 139
temperature stations in the UK. See Section 2 of the SI for
further information on the meteorological simulation and
validation.
The regional chemistry-transport model CMAQ19 is applied
to simulate air quality in Europe at a resolution of 40.5 km, with
a nested 13.5 km grid for the UK. The global chemistrytransport model GEOS-Chem20 is applied for 2005 to provide
boundary conditions to the CMAQ 40.5 km European domain.
3. RESULTS AND DISCUSSION
3.1. Model Evaluation. A set of statistical measures as
recommended by U.S. EPA is estimated30 including index of
agreement. An index of agreement of 1 indicates perfect
agreement between the model and observations, while 0
indicates no agreement. On average, the WRF model achieved
0.83 indices of agreement for both wind speed and temperature. The simulated mean wind speed has a bias of +18% and
temperature (in °C) −6% relative to observations. Further
statistical parameters are given in Table 4 in the SI.
Average indices of agreement for O3, NO2, PM10, and PM2.5
are 0.63, 0.53, 0.5, and 0.7, respectively. Other model evaluation
metrics are shown in Tables 5 and 6 of the SI. The annual mean
PM2.5 modeling bias for all stations is −28% or −20% excluding
the roadside station, where the greatest bias is −52% and the
smallest is −9%. We note that the average bias for PM10 in the
UK is −65%, likely due to incomplete representation of coarse
PM in our PM2.5-focused setup. We represent the uncertainty
in our CMAQ results for PM2.5 as having a nominal bias of
−28% with an uncertainty range of −65% to −9%, where the
lower bound has been extended to capture the mean bias in
PM10. This assumed bias is due to a combination of emissions
and atmospheric modeling uncertainty. (See Section 3 in the SI
for further discussion on emissions uncertainty and assessment
of modeling biases.)
3.2. PM2.5 Impacts. Figure 1 depicts the annual average
ground-level PM2.5 concentration due to different combustion
sources. A gradient from northwest to southeast is observed
which is consistent with the finding in Stedman et al.8 Road
transport contributes the highest proportion of the annual
average ground-level PM2.5 concentration among all sectors,
especially in southeast England. The population-weighted
PM2.5 concentration attributable to road transport in the UK
is 0.75 μg/m3. Figure 2 illustrates annual average ground-level
soot (black carbon (BC)) and nitrogen dioxide (NO2)
concentrations attributable to road transport. The groundB
Policy Analysis
Figure 3. Annual average (a) BC concentration due to UK other
transport and (b) SO4 concentration due to UK power generation
emissions.
representing approximately half of total national SOx emissions.
Figure 3b depicts the annual average ground-level sulfate
concentration due to power generation. The Northern East
Midlands region has a ground-level sulfate perturbation of up to
1 μg/m3. This is likely attributable to the five major (>1900
MW capacity) power plants located in that region. In part due
to the southwesterly prevailing winds in southeast England, the
population-weighted ground-level PM2.5 concentration in
London due to power generation is 0.38 μg/m3a third of
road transport’s impact.
Finally, combustion emissions from commercial buildings,
institutional, residential, agriculture, and industries together
contribute 0.26 μg/m3 to the population-weighted PM2.5
concentration in the UK. Impacts are approximately even
throughout southern and central England.
3.3. Health Impacts (Nominal Estimates). Applying the
central CRF (a 1% increase in risk of premature mortality per
μg/m3 of long-term exposure to PM2.5) to our CMAQ results
we calculate “nominal” estimates of early deaths per year
attributable to each sector, which are shown in Table 1. A total
Figure 1. Annual average PM2.5 concentration due to combustion
emissions from (a) power generation; (b) commercial, institutional,
residential, and agricultural sources; (c) industry; (d) road transport;
(e) other transport; and (f) all UK combustion sources.
Figure 2. Annual average (a) BC and (b) NO2 concentration due to
UK road transport emissions.
Table 1. Central Estimates for Early Deaths Per Year in the
UK by Combustion Sector Calculated Using the High
Resolution Modeling Domaina
level [BC] and [NO2] perturbation due to road transport has
local peaks in cities due to the localized nature of BC and NOx
emissions and their direct (i.e., non-secondary) impacts. By
contrast, the total (primary + secondary) PM2.5 impacts of road
transport are relatively diffuse (Figure 1a). Population-weighted
[PM2.5] due to road transport in London is 1.21 μg/m3, which
is 1.6 times higher than the UK average for populationweighted [PM2.5] due to road transport.
Other transport is the second highest contributor to
population-weighted annual average ground-level [PM2.5]
(closely followed by power generation). According to the
NAEI, other transport produces 21% (0.02 Tg (−20 to +30%))
of UK annual primary PM2.5 emissions, ahead of other sectors.
Figure 3a shows the annual average ground-level [BC] due to
other transport. It illustrates the local peaks of BC at (marine)
ports and airports. Other transport contributes 0.42 μg/m3 to
population-weighted [PM2.5] in the UK. It is calculated that the
London population-weighted [PM2.5] attributable to other
transport emissions is 0.51 μg/m3. This is partly associated with
the London airports, including Heathrow, Luton, Gatwick,
Stansted, and London City.
Combustion emissions from power generation result in an
average population-weighted PM2.5 concentration of 0.4 μg/m3
in the UK. Among different PM2.5 species due to power
generation, sulfate aerosol accounts for 62% of the total
population-weighted PM2.5. According to the NAEI, power
generation is responsible for 0.29 Tg (±4%) of SOx emissions,
sector
power generation
commercial, institutional,
residential and agriculture
industry
road transport
other transport
all UK combustion
nominal early
deaths/year
(UK)
corrected central estimate
and uncertainty (90% CI)
1700
1100
2500 (1400−3800)
1600 (850−2400)
560
3300
1800
9000
830 (440−1200)
4900 (2600−7200)
2600 (1400−4000)
13,000 (6900−20,000)
a
Note that results from simulations of individual sectors do not sum
exactly to results from a simulation of all UK combustion sectors due
to nonlinearities. (Nonlinearities result in a 6% discrepancy in early
deaths when comparing the sum of sector simulations to the all-sector
simulation results.)
of 9000 UK premature mortalities are estimated to be
attributable to UK combustion emissions, of which 3300,
1800, and 1700 deaths per year are due to road transport, other
transport, and power generation emissions, respectively.
We note that the road transport estimate in particular is
likely to be an underestimate, as the peaks in roadside PM2.5
may not be accurately represented due to our model resolution.
For example, while CMAQ underestimates [PM2.5] by 20% on
average at the non-roadside PM monitoring stations, it
C
Policy Analysis
underestimates the average PM2.5 concentration by 52% at the
London Marylebone Road station (located 1 m from the curb).
We use ArcGIS to overlay CMAQ gridded results for each
sector onto UK administrative regions. The nominal estimate
for number of premature deaths caused by different combustion
sources in different UK administrative districts is shown in
Section 5 of the SI, along with the definition of districts. We
find that power generation has greater health impacts than
other transport in South Yorkshire districts and Yorkshire and
the Humber region; however, compared to power generation,
other transport causes more combustion emissions-attributable
premature mortalities in the South West and East districts,
Merseyside district, and London.
Road transport contributes 27−50% (36% on average) of the
total combustion emissions-attributable premature mortalities
across the different districts. In particular, the results show that
road transport causes 660 premature mortalities in Greater
London, representing half of total premature mortalities in the
city associated with all combustion emissions.
For London, we calculate that the premature mortalities
associated with the other transport and power generation are
280 and 200, respectively. The higher health impacts due to
other transport compared to power generation may be
associated with airports around Londonespecially Heathrow
to the west and Gatwick to the southand the prevailing
southwesterly wind. This means that power generation
emissions, which are predominantly to the northeast of major
population centers and have higher effective emissions heights,
are less damaging on a per unit emission basis to UK public
health.
3.4. Uncertainty Estimation. We estimate the uncertainty
in the premature mortality calculations. The uncertainty in the
CRF is accounted for with a triangular probability distribution
of multipliers with (low, nominal, high) values of (0.355, 1.06,
1.81),23 where the low, nominal, and high values correspond to
the vertices of the distribution function. The bias and
uncertainty in modeling PM concentrations is represented by
a triangular distribution of multipliers with (low, nominal, high)
values of (1.09, 1.25, 1.65). A uniform distribution with range
(1, 1.06) is selected to account for the uncertainty due to the
nonlinearities when comparing the sum of sector simulations to
the all-sector simulation results. The ∼10% probability of no
causal link between PM2.5 exposure and premature mortality
has not been accounted for quantitatively. Uncertainty ranges
and corrected central estimates accounting for biases in
modeling are shown in Table 1, where ranges are 90%
confidence intervals. For example, the corrected central
estimate for premature mortalities attributable to road transport
emissions is 4900 per year with a range of 2600−7200, which
can be compared to the nominal estimate of 3300. Estimates for
other sectors are given in Table 1.
We note that a potentially significant unquantified
uncertainty is the differential toxicity among PM species.
Expert committees have concluded that there is currently no
strong basis for an alternative to the assumption of equal
toxicity among PM species.30 However, BC is likely more toxic
than other PM constituents,31 which indicates that the health
impact of road transport is likely to be further underestimated.
3.5. Transboundary Impacts. Understanding transboundary air pollution is of importance when considering environmental policy measures at an EU level. Figure 4a illustrates the
annual average PM2.5 concentration associated with combustion
emissions from non-UK EU sources. We find that the PM2.5
Figure 4. Impact of non-UK EU combustion emissions on the UK
expressed as (a) an absolute perturbation, and (b) the percentage of
total UK PM2.5 contributed by non-UK EU combustion emissions.
perturbation due to non-UK EU combustion emissions is
approximately 2 μg/m3 in the southeast of the UK and ∼1 μg/
m3 in the Midlands. This result is consistent with the finding of
Malcolm et al.32
The population-weighted PM2.5 concentration in London
due to non-UK EU combustion emissions is 1.17 μg/m3, which
is approximately equal to the impact of UK road transport
emissions (1.21 μg/m3). We estimate that 4100 (nominal
estimate) early deaths are associated with non-UK EU
combustion emissions, of which 650 are in London as shown
in Table 2. (A corrected central estimate and uncertainty range
is also given in Table 2.)
Figure 4b shows the percentage of total UK PM 2.5
contributed by non-UK EU combustion emissions. (Here
“total” is the sum of PM2.5 caused by UK plus non-UK EU
combustion emissions.) It can be seen that the minimum value
is 30% in the Midlands and the Central Belt in Scotland, while
in the Highlands >70% of [PM2.5] is due to non-UK EU
emissions. Our results show that [PM2.5] attributable to nonUK EU combustion emissions accounts for approximately 40%
of the total along parts of the south and east coasts of England.
The PM2.5 concentration at monitoring stations at Rochester,
Barcombe Mills, Southampton, Yarner Wood, and Stoke
Ferryanalyzed in Malcolm et al.also indicates that ∼40%
of total [PM2.5] comes from outside UK.
Table 2 depicts the nominal and corrected central estimates
(with uncertainty ranges) for early deaths per year in the UK
and the rest of the EU. Results for the rest of the EU are based
on the lower resolution (40.5 km) CMAQ results. Non-UK EU
combustion emissions result in 4100 (nominal estimate)
premature deaths in the UK per year, while UK combustion
emissions account for 8500 early deaths in the UK. This
indicates that transboundary pollution accounts for one-third of
the combustion emissions-attributable deaths per year in the
UK. Conversely, UK combustion emissions cause 3100
premature mortalities per year in other EU member states, or
2% of the total 130,000 early deaths per year in the non-UK
EU. As non-UK EU emissions are higher than the UK alone,
this implies that on a per unit emission basis, the UK “exports”
more public health damage to the rest of the EU than it
“imports”. This is consistent with prevailing southwesterly/
westerly winds.
3.6. Implications for Policy. The road transport sector is
found to be the major contributor to PM2.5 exposure in the UK,
and the resulting premature mortalities are comparable to the
2946 deaths due to road accidents in 2007,33 indicating that the
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Policy Analysis
Table 2. Central Estimates for Early Deaths Per Year in the UK and the Rest of the EU Calculated Using Results from the LowResolution Domain for Non-UK Deaths and the High-Resolution Domain for UK Deathsa
emissions sources removed
all UK combustion
all non-UK EU combustion
total
a
UK early deaths/year
non-UK EU early deaths/year
greater London early deaths/year
8460
[12,000 (6800−19,000)]
4100
[6000 (3200−9000)]
12,560
3100
[4500 (2400−6700)]
130,000
[190,000 (100,000−290,000)]
133,100
1500
[2200 (1100, 3300)]
650
[960 (530−1400)]
2150
Corrected central estimates and uncertainty ranges are shown in italics.
Notes
public health impacts of road transport are likely to be 50%
greater than fatal accidents as measured by attributable
premature mortalities. (The number of deaths on UK roads
decreased to 1850 in 2010.) We note that an air quality-related
mortality is not equivalent to a fatal road accident in terms of
life years lost on average. For example, approximately half of
those who died on UK roads in 2007 were under 40, implying a
loss of life of ∼35 life years per mortality, compared to the ∼12
life years lost per air quality mortality estimated by COMEAP.
This means that road accidents are still likely to result in a
greater loss of life years than road transport emissions.
Approximately one-sixth of PM2.5 exposure attributable to
transport (as a whole) is BC (see Tables 7 and 8 in the SI).
This can be compared to 1−2% for other sectors and is
indicative of the extent to which road transport has localized
impacts due to the positive correlation between road transport
emissions and population density. On the other hand, sulfate
impacts of road transport represent 1% of the sector’s total
PM2.5 impact, which can be compared to figures of 10% for
industry to 62% for power generation. This is consistent with
the low sulfur fuel used in road transport in the UK and the
high sulfur coal-fired power stations in use. Taken together,
these findings suggest further efforts to reduce UK power
station SOx emissions should be assessed for their costs and
benefits, while for road transport the planned reductions in
allowable primary PM emissions may have significant health
benefits.
In terms of economic impacts, we estimate that combustion
emissions cost the UK £6−62bn/year. This corresponds to
0.4−3.5% of UK gross domestic product in 2007. The bounds
correspond to medians of typical EU and U.S. approaches to
monetizing early deaths (see Section 6 of the SI).
The extent of transboundary pollution between the UK and
other EU member states can be illustrated by noting that (i)
one-third of premature mortalities in the UK caused by
combustion emissions are due to emissions from other EU
member states, and (ii) UK combustion emissions cause onethird again as many early deaths in the rest of the EU as they do
in the UK. These results indicate that further policy measures
should be coordinated at an EU-level because of the strength of
the transboundary component of PM pollution, and that the
EU as a whole is responsible for air quality in any given
member state.
■
The authors declare no competing financial interest.
■
ACKNOWLEDGMENTS
This work was started under the Energy Efficient Cities
initiative at the University of Cambridge, funded by EPSRC.
We thank EPSRC for initial funding and MIT for supporting
the conclusion of the work.
■
ASSOCIATED CONTENT
* Supporting Information
S
Further discussion, analyses, and results. This material is
available free of charge via the Internet at http://pubs.acs.org.
■
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Corresponding Author
*Phone: +1 617-452-2550; e-mail: [email protected].
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