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Author's personal copy
Environmental Pollution 150 (2007) 150e165
www.elsevier.com/locate/envpol
Review
Global fate of POPs: Current and future research directions
Rainer Lohmann a,*, Knut Breivik b,c, Jordi Dachs d, Derek Muir e
a
Graduate School of Oceanography, University of Rhode Island, Narragansett, RI 02882-1197, USA
b
Norwegian Institute for Air Research, PO Box 100, NO-2027 Kjeller, Norway
c
University of Oslo, Department of Chemistry, PO Box 1033, NO-0315 Oslo, Norway
d
Department of Environmental Chemistry, Institute of Chemical and Environmental Research (IIQAB-CSIC), Jordi Girona 18e26, Barcelona 08034, Spain
e
Aquatic Ecosystem Protection Research Division, Environment Canada, 867 Lakeshore Road, Burlington, ON L7R4A6, Canada
Received 30 March 2007; received in revised form 5 June 2007; accepted 8 June 2007
Future studies into the global fate of POPs will need to pay more attention to the various biogeochemical
and anthropogenic cycles to better understand emissions, transport and sinks.
Abstract
For legacy and emerging persistent organic pollutants (POPs), surprisingly little is still known in quantitative terms about their global sources
and emissions. Atmospheric transport has been identified as the key global dispersal mechanism for most legacy POPs. In contrast, transport by
ocean currents may prove to be the main transport route for many polar, emerging POPs. This is linked to the POPs’ intrinsic physico-chemical
properties, as exemplified by the different fate of hexachlorocyclohexanes in the Arctic. Similarly, our current understanding of POPs’ global
transport and fate remains sketchy. The importance of organic carbon and global temperature differences have been accepted as key drivers of
POPs’ global distribution. However, future research will need to understand the various biogeochemical and geophysical cycles under anthropogenic pressures to be able to understand and predict the global fate of POPs accurately.
Ó 2007 Elsevier Ltd. All rights reserved.
Keywords: Persistent organic pollutant; POP; Sources; Fate; trends
1. Introduction
Persistent organic pollutants (POPs) constitute a diverse
group of organic substances, which are toxic, persistent, bioaccumulative and prone to long-range transport. They have
different intrinsic physical-chemical properties, which dictate
their environmental behavior (see Wania, 2003, 2006;
Fig. 1). Equally important, their sources and pathways for release into the global environment vary as well. POPs can be
classified as flyers, multi-, single hoppers and swimmers
(Fig. 1). This assignment was based on the major modes of
chemical transport behavior on a global scale according to
* Corresponding author: Tel.: þ1 401 874 6612; fax: þ1 401 874 6811.
E-mail address: [email protected] (R. Lohmann).
0269-7491/$ - see front matter Ó 2007 Elsevier Ltd. All rights reserved.
doi:10.1016/j.envpol.2007.06.051
their specific partitioning property combinations by Wania
(2003). The assessment was based on the results of Arctic
Contamination Potential (ACP) for emissions to air over
a 10-year period using calculations for perfectly persistent
chemicals (Wania, 2003). Most classical POPs (e.g., polychlorinated biphenyls (PCBs), DDT, chlorobenzenes and organochlorine pesticides (OCPs) such as chlordane and
toxaphene) can be classified as multi-hoppers (Wania, 2003,
2006) (see Fig. 1A). In contrast to these multi-media chemicals, the hexachlorocyclohexanes (HCHs) are in the ‘‘swimmers’’ category. Fig. 1B shows the approximate positions of
a range of chemicals in commerce based on their predicted
or measured log KAW and log KOA.
The UNEP Stockholm Convention (UNEP, 2001) currently
regulates the so-called ‘‘dirty dozen’’ or ‘legacy’ POPs (see
Table 1). Additional substances and substance groups are
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R. Lohmann et al. / Environmental Pollution 150 (2007) 150e165
B
A
3
3
lo
3
Polyfluoro alcohols
PBDE/HxBCD
2
Bromo-benzenes/cyclohexanes
Cyclic siloxanes
g
CBz
2
1
HCHs
DDTs
1
=
2
W
Cyclic siloxanes
KO
PCBs
8
Polyfluoro alcohols
8:2 FTOH
NEtFOSE
NEtFOSA NMeFOSE
0
0
C10-C1chlorinated paraffins1
Haloalkyl phosphates
Triaryl phosphates
0
lo
-1
-1
KO
chlorobenzenes
g
“Classic” Arctic POPs
Br-benz/cyclohex
3
-2
=
-1
W
log KAW
151
PCBs
-2
-3
-3
DDT
-4
3
4
5
6
7
fliers
8
9
-3
Haloalkyl
phosphates
TeBDE
HxBCD
-4
DDE
HCHs
10
11
PeBDE
-4
Chlorinated paraffins
Perfluor
-5 o acids
12
4
3 PFOS,
log KOAPFCAs
multiple hoppers
-2
phosphates
5
swimmers
6
7
8
9
10
-5
11
12
single hoppers
Fig. 1. Major modes of transport of perfectly persistent, hypothetical chemicals defined by their partitioning properties log KAW and log KOA, when calculated with
the Globo-POP model (Wania, 2003) assuming 10 years of steady emissions into air. Graphics modified from Wania (2006), with copyright by the American
Chemical Society). (A) Legacy organochlorine compounds. (B) Some chemicals in current use. (The arrow for perfluoro acids illustrates that these degradation
products of the polyfluoro alcohols have log KAW 5 and log KOA 3, and are thus off scale).
covered by the United Nations/Economic Council for Europe
Protocol on POPs (UNECE, 1998) (see lower part of Table 1).
It is convenient to discriminate between chemicals that are intentionally produced and chemicals that are formed as accidental by-products of various combustion processes. In the
former group, there is a wide range of OCPs and some industrial chemicals, whilst the latter group captures substances like
polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/
Fs) and polycyclic aromatic hydrocarbons (PAHs). PAHs are
only recognized as POPs under the Aarhus Protocol (UNECE,
1998). However, recent studies have highlighted that many industrially produced compounds are also emitted during
combustion processes. This will likely reflect a combination
of de novo formation of POPs such as PCBs, PCNaphthalenes,
hexachlorobenzene (HCB) (e.g., Bailey, 2001; Lee et al.,
2005) and the volatilization of intentionally produced compounds such as polybrominated diphenylethers (PBDEs) (Farrar et al., 2004).
Both international agreements mentioned earlier include
mechanisms for adding new substances and substance groups
if certain specified criteria are met (UNECE, 1998; UNEP,
2001). Several ‘‘new’’ chemicals of emerging environmental
concern are currently being scrutinized for possible POP-like
behavior. These potentially new POPs (‘‘emerging’’ POPs)
Table 1
Legacy POPs controlled under the Stockholm Convention (UNEP, 2001) and the Aarhus Protocol on POPs (UNECE, 1998), including available estimates of the
global cumulative production and usage of some intentionally produced POPs or annual emissions (modified after Table 2.1 in de Wit et al., 2004)
Chemical
Key source
Aldrin
Chlordane
Dieldrin
Endrin
Heptachlor
Mirex
Toxaphene
Insecticide
Insecticide
Insecticide
Insecticide
Insecticide
Insecticide, flame retardant
Insecticide
DDT
PCB
HCB
PCDD/Fs
Hexabromobiphenyla
HCHsa
Technical HCH
Lindane
PAHsa,b
Chlordeconea
Insecticide
Industrial chemical
Miscellaneous
By-product
Flame retardant
Insecticide
a
b
By-product
Insecticide
Production period
or reference year(s)
for emission estimates
Cumulative global
production/usage, kt)
or annual emissions
Reference
1950e1993
1330 kt
1940sepresent
1930e1993
w1995
w1995
4500 kt
1326 kt
w23 t/year
9.9 kg TEQ/year
Voldner and Li (1993) see also
Li and Macdonald (2005)
Li and Macdonald (2005)
Breivik et al. (2007)
Bailey (2001)
UNEP (1999); Fiedler (2003)
1948e1997
1950e1993
1966e1969
10000 kt
720 kt
5000 t/year
Li (1999)
Voldner and Li (1995)
Suess (1976)
Only regulated under the Aarhus Protocol on POPs (UNECE, 1998).
Data refers to B[a]P only (Suess, 1976).
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R. Lohmann et al. / Environmental Pollution 150 (2007) 150e165
have in common that they are generally intentionally produced
chemicals. Examples are short chain chlorinated paraffins,
hexachlorobutadiene, pentachlorobenzene, ‘‘penta’’ and
‘‘octa’’ BDEs, and perfluorinated octyl sulfonamides and sulfonates (PFOS and related chemicals) which, in early 2007,
were being considered for future inclusion in the UNEP
POPs and UNECE protocols. Recent screening of chemicals
in commerce has revealed that many other intentionally produced chemicals share similar (predicted) physico-chemical
characteristics to those currently under consideration. However, they have generally not been measured in the global environment (Muir and Howard, 2006). Fig. 1B includes some
examples. In fact, while we know a few legacy POPs rather
well, the presence of most anthropogenic organic chemicals
in the different environmental compartments is unknown.
So far, environmental chemistry has focused on multiple
and single hoppers, whereas many ‘‘emerging POPs’’ are polar
compounds or ‘‘swimmers’’ (e.g., PFOS, brominated cyclohexanes; see Fig. 1B). Hoppers are relatively easy to determine by gas chromatography (GC) analysis of atmosphere,
vegetation, soil and sediments samples. In contrast, the sampling and analysis of water poses many logistical and analytical challenges (e.g., Lohmann et al., 2004). For polar and
non-volatile compounds analysis will rely more on liquid
chromatography, particularly for ionic chemicals. More importantly, while the POPs protocol was based on the detection of
specific chemicals in the field, this will have to change in the
future. Recent studies highlight how we can identify new
POPs (e.g., Muir and Howard, 2006), and suggest ways of prioritizing them (e.g., Arnot et al., 2006). In the following, we
review the state-of-the-art regarding the global fate of POPs.
In particular, we try to point out the lessons we have learned
from studying legacy POPs, and the implications this might
have for addressing remaining challenges for legacy POPs
and new issues for intentionally produced chemicals in commerce which may have the characteristics of POPs.
Ideally, we would like to draw up a ‘‘global mass balance’’,
which combines knowledge of source emission rates with the
quantification of environmental reservoirs and final sink
fluxes. However, coming up with accurate emission rates remains a challenge, while we are still improving our understanding of the global distribution of POPs. Easier tasks are
to account for inventories of POPs (e.g., their presence in soils
or sediments), but we have more difficulties in estimating historical emissions, removal fluxes through reactions and the
presence of POPs in remote locations, such as the deep oceans.
2. Emissions of POPs
2.1. Global emissions
Information on the sources and emissions of POPs may be
among the least certain aspects of the overall fate of these
compounds in the global environment (Wania and Mackay,
1996; Vallack et al., 1998; Jones and de Voogt, 1999), although a number of regional and global emission inventories
for POPs and emerging POPs have become available recently
(Voldner and Li, 1993; UNEP, 1999; Macdonald et al., 2000;
Li and Bidleman, 2003; Breivik et al., 2004; de Wit et al.,
2004; Fiedler, 2007). We have listed best available estimates
on the global historical production or consumption of intentionally produced legacy POPs, or annual global emissions
in the case of PCDD/Fs, PAHs and HCB (Table 1).
Several studies have highlighted the difficulty in obtaining
the relevant and reliable information on production and use of
POPs (e.g., Breivik et al., 2006). Specifically, such data may
be confidential, proprietary or not recorded on a routine basis
by many countries (Voldner and Li, 1995). Even for legacy
POPs, there is a general lack of their global production and
consumption data (Table 1). Against this background, studies
on the usage and emissions of various OCPs by Li and coworkers stand out as landmarks. Specifically, much emphasis
has been devoted to understanding and predicting the global
sources and emissions of technical HCHs and its key constituents (Voldner and Li, 1995; Li et al., 2000, 2002, 2003, 2005,
2006; Li, 2001; Li and Bidleman, 2003; Li and Li, 2004; Li
and Macdonald, 2005). These studies have been essential in
terms of facilitating consecutive model studies to predict the
global transport and fate of a-HCH, the major by-product of
technical HCH (Strand and Hov, 1996; Wania and Mackay,
1999; Wania et al., 1999; Koziol and Pudykiewicz, 2001;
Toose et al., 2004).
More recently, a related attempt was made by Breivik et al.
(2002a,b) to determine the global historical emissions of 22
different PCB congeners from 1930 to the year 2000, using
a dynamic mass balance (or ‘‘cradle-to-emissions’’) approach.
The latter estimates have recently been updated in an attempt
to improve certain temporal features of the methodology and
to present scenarios for future emissions (Breivik et al., 2007).
There have been various additional attempts in the literature
to quantify global emissions of other POPs (e.g., Suess, 1976;
Bailey, 2001; Fiedler, 2003). Although such information may
be suitable for mass balance studies (e.g., Brzuzy and Hites,
1996) and to identify key source categories on a global scale
they typically lack spatial and/or temporal resolution, which
again limits their use by spatially and temporally resolved
global fate models.
Recent developments in the registration and regulation of
chemicals in Europe and the US may also improve our ability
to estimate emissions. In the EU, the REACH regulation (Registration, Evaluation, Authorisation and Restriction of Chemicals) will require the registration of substances manufactured
or imported in volumes starting at 1 t/year and risk assessments of existing substances marketed in volumes at or above
10 t/year (European Commission, 2006). In 2006, the US
EPA’s Inventory Update Rule will include consumer end
uses of chemicals for substances produced or imported at
>136 t/year although the threshold for reporting will be raised
to 11.4 t/year (USEPA, 2006).
2.2. Regional emissions
Although the available information on a global scale may be
limited, there are several regional emission inventories
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available. For Europe, Duiser and Veldt presented a pioneering
study addressing selected PAHs, PCBs, lindane and HCB for the
reference year 1982 (Duiser and Veldt, 1989) with more compounds being included for the reference year 1990 (Berdowski
et al., 1997). In its most recent update, estimates of both current
and future emission scenarios for both legacy POPs and emerging POPs were presented (Denier van der Gon et al., 2005). Additional studies include the PCDD/F emission inventories
presented for various European countries (Quass et al., 2000,
2004; Pulles et al., 2006). Furthermore, there have been studies
with emphasis on establishing temporal trends in emissions
across Europe from 1970 to the mid-1990s for OCPs, HCB
and PCDD/Fs (Breivik et al., 1999; Pacyna et al., 2003), thus facilitating an analysis of the environmental response to changes
in emissions within this region (the response of the Arctic to
changing emissions is detailed below).
A number of regional inventories for various OCPs have
been published by Li and co-workers. Specifically, data on
pesticide usage and emissions have become available in the
case of toxaphene in the United States (Li, 2001; Li et al.,
2001a), lindane usage in Canada (Li et al., 2004b), HCH
and DDT usage in the former Soviet Union (Li et al., 2005,
2006) as well as HCH usage in China (Li et al., 2001b).
More recently, a detailed emission inventory for PAHs in
China was published by Tao and co-workers (Xu et al.,
2006; Zhang et al., 2007).
It is anticipated that the two international agreements mentioned above will motivate further development of emission
inventories for POPs. There are significant ongoing efforts to
further develop relevant control strategies under the UNEP
Stockholm Convention (UNEP, 1999/2001; Fiedler, 2007).
The parties to the Stockholm Convention are obliged to develop National Implementation Plans and, as elements of these
plans to draw up POPs emission inventories. For example,
a so-called toolkit has been developed to assist countries in
quantifying dioxin emissions in a consistent manner (UNEP,
2005).
Within Europe and North America, parties of the 1998 UNECE POPs Protocol are additionally requested to submit their
national emission data (UNECE, 2002), which are summarized in annual reports and made available though the internet
(e.g., Vestreng et al., 2006). Most data are available for PCDD/
Fs and PAHs, presumably in part because existing emission inventory methodologies for the better characterized classical air
pollutants (SO2, NOx) could easily be modified to include
combustion related by-products (Breivik et al., 2006). Finally,
considerable efforts are currently being devoted to the compilation of PCB emission inventories in Eastern Europe.
2.3. Primary versus secondary sources
The contemporary environmental burden of POPs reflects
both past and current primary emissions (Wania and Mackay,
1993, 1996). Secondary emissions are due to the POPs’ environmental persistence and potential for reversible atmospheric deposition to and from aquatic and terrestrial environments (i.e.,
hopping, see Fig. 1). Even for industrially produced legacy
153
POPs whose production ended a long time ago, primary sources
can contribute strongly to current emissions (e.g., Kohler et al.,
2005). This has obvious implications for evaluating potential
control strategies as well as understanding POPs’ emissions
on a global scale (e.g. Bidleman and Falconer, 1999).
HCB offers a prime example of a legacy POP whose
contemporary atmospheric level is significantly controlled by
environmental re-emissions (Bailey, 2001; Jaward et al.,
2004c; Barber et al., 2005; Su et al., 2006). In contrast to
HCB, atmospheric levels of higher molecular PAHs are presumably controlled by current primary atmospheric emissions
because of the PAHs’ strong sorption to atmospheric particles
(e.g., Dachs and Eisenreich, 2000; Lohmann and Lammel,
2004), i.e. PAHs act as ‘‘single hoppers’’ (Fig. 1).
For other POPs, such as PCBs, there has been an extensive debate whether contemporary atmospheric levels are mainly due to
environmental cycling (e.g., Larsson, 1985; Harrad et al., 1994;
Jeremiason et al., 1994), or due to primary emissions (e.g.,
Jaward et al., 2004c; Robson and Harrad, 2004; Hung et al.,
2005b). An early study in the United Kingdom suggested that
the major source of PCBs to the atmosphere was volatilization
from soils (Harrad et al., 1994). However, a more recent European study observed elevated levels in urban air and attributed
these to continuing diffusive atmospheric emissions in densely
populated areas (Jaward et al., 2004c). Although this strongly
suggests that there still are continuing emissions in urban regions, it does not really confirm whether these are due to primary
(e.g. volatilization from electrical equipment or building materials) or secondary sources (re-emissions from contaminated environmental hot spots) (see also Hsu et al., 2003; Uraki et al.,
2004). However, very recent results from Birmingham (UK)
strongly suggest that indoor air is of greater importance than
volatilization from soils in controlling atmospheric levels in
this region (Jamshidi et al., 2007).
As ‘‘multi-hopping’’ (e.g., Wania and Mackay, 1996) could
strongly affect contemporary atmospheric levels, numerous approaches have been used to evaluate contemporary fluxes from
contaminated environmental reservoirs (Bidleman and Falconer, 1999). These include interpreting chiral compounds by
Bidleman and others (e.g., Ridal et al., 1997; Finizio et al.,
1998; Bidleman and Falconer, 1999; Robson and Harrad,
2004; Jamshidi et al., 2007), the evaluation of fugacity ratios/
fractions (e.g., Jantunen and Bidleman, 1995; Kurt-Karakus
et al., 2006) and the use of multimedia fate and transport models
(e.g., Scheringer et al., 2000; Wegmann et al., 2004; Hung et al.,
2005b). A better insight into the relative significance of primary
and secondary sources through a combination of complementary field, laboratory and modeling efforts is still needed. An interesting implication is the contrast between rural agricultural
areas as sources of legacy POPs and new pesticides, while industrial-urban areas are sources of PCDD/Fs, PCBs and other
consumer compounds linked to population density.
3. Emission and exposure trends in the Arctic
Long term time trends in air, water and biota can be very
useful in inferring global fate and cycling of POPs. However,
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there are relatively few long-term datasets for temporal trends
of legacy POPs that can be used as ‘‘global’’ indicators of
trends. To be a good global indicator the dataset would need
to be in a remote environment, removed from local source influences. Thus strong temporal trend datasets for POPs in biota
in the Baltic (Olsson et al., 2000) and the Great Lakes (EC and
USEPA, 2005) may not be the best for global purposes. Global
background concentrations in air have been determined for
12 years at Alert (Northern Ellesmere Is, Canada) (Hung
et al., 2005a). This is an extremely detailed dataset, however,
the time span covers only a part of the post-1980s bans on
OCPs and open uses of PCBs in most circumpolar countries
(de Wit et al., 2004).
3.1. Time trends of HCHs
Time trends of POPs in seawater are also very limited and
again the most detailed datasets are from the Arctic Ocean (de
Wit et al., 2004). Li and Macdonald (2005) have reviewed historical global usage and emissions for OCPs, and noted that
the most comprehensive data existed for HCHs. Air measurements of a-HCH in the Arctic have been made as far back as
the late 1970s. They show a declining trend (Fig. 2) during the
1980s and 1990s, which paralleled the estimated global emissions of a-HCH (Li et al., 2000).
Based on a box model of HCH in the Arctic environment,
Li et al. (2004a) concluded that ocean currents became the major pathway for a-HCH to enter the Arctic Ocean after the
early 1990s. At that time decreasing atmospheric a-HCH
emissions were superseded by the historical burden of aHCH in upper ocean waters, accumulated from over 40 years
of use of technical HCH. Temporal trends for HCH isomers in
seawater are much more limited than air measurements. However, data assembled by Li and Macdonald (2005) suggests
that a-HCH concentrations in North American Arctic waters
(Beaufort Sea and Canadian archipelago) declined during the
1990s while b-HCH peaked in the mid-1990s. These time
trends reflect different pathways for the two isomers from
the major source region, China (Li et al., 2002). Although
emitted along with a-HCH (from China, the major source region), b-HCH has 20 times lower Henry’s Law Constant
(HLC) and therefore was deposited into the North Pacific because of more efficient scavenging by precipitation and air to
sea deposition by gas exchange.
3.2. Use of Arctic biota as monitoring tools
Since the HCH isomers bioaccumulate in the marine food
web, their time trends can be followed using top predators
such as seals and seabirds. Tanabe et al. (Dempster et al.,
1994) noted that the proportion of b-HCH increased over
time in northern fur seals (Callorhinus ursinus) inhabiting
the western North Pacific Ocean. However, a distinctive trend
indicative of separate pathways for a- and b-HCH was not apparent in their study. Further from the source region, distinctive trends for a- and b-HCH in Canadian Arctic ringed
seals (Phoca hispida) have been found in populations from
the Lancaster Sound region (Fig. 3).
a-HCH has declined about 3-fold since the early to mid1980s in these seals. This is in good agreement with trends
seen for total HCHs in northern fur seals in the 1980s and
1990s (Kajiwara et al., 2004). However, b-HCH concentrations in the same animals reached a maximum in the mid1990s and have since declined. Similar trends are evident
for a- and b-HCH in seabird eggs from Lancaster Sound
where a distinctive increase of b-HCH was evident during
the 1990s (Braune et al., 2001). Overall, the trends in Arctic
seals do not follow the predicted global emissions of the two
isomers (Fig. 3) particularly for b-HCH (Li et al., 2003).
The temporal trend is best explained by the pathways described by Li et al. (2002) in which b-HCH residues, which
are confined to upper ocean layers due to its relatively high
water solubility, arrived in Lancaster Sound mainly via ocean
transport through the Bering Strait and eastward in the Beaufort Sea through the Canadian archipelago.
3.3. Other POPs in the Arctic
Fig. 2. Global emissions of a-HCH (Li et al., 2000) and its mean concentrations in Arctic air from 1979 to 1996 assembled from published data (de
Wit et al., 2004). This figure is reproduced from DeWit et al. with permission
of the AMAP (Oslo No).
SDDT (sum of o,p0 - and p,p0 -DDT, DDE and DDD) concentrations in ringed seals, have declined steadily since the
mid-1970s in Lancaster Sound (Fig. 3). Li and Li (2004)
have estimated the global emissions from agriculture from
1947 to 2000 and results for 1960e2000 are plotted in
Fig. 3. The decline in emissions post-1970 is paralleled very
well in ringed seals and a plot of emission vs SDDT for the
years in which seals were sampled gives an r2 ¼ 0.93
(N ¼ 6, P ¼ 0.002). Similarly, the declines of S10PCB (sum
of CB 28, 31, 52, 101, 118, 105, 153, 138, 156, 180) in seals
parallels reasonably well the decline in global PCB emissions
reported by Breivik et al. (2002a,b).
The distinctive differences in trends between DDT, PCBs
and b-HCH in the ringed seals from a remote background
site illustrate differences between ‘‘hoppers’’ (DDT, PCBs)
and ‘‘swimmers’’ (HCHs) (see Fig. 1). In addition to generally
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R. Lohmann et al. / Environmental Pollution 150 (2007) 150e165
400
α-HCH
α-HCH
emissions (kT)
α-HCH
300
250
50
β-HCH
β-HCH emissions (kT)
155
20
β-HCH
200
40
16
150
30
12
100
20
8
50
10
4
200
ng/g lipid wt
0
1960
0
1970
1980
500
1990
2000
2010
Σ10PCB
PCB emissions (T)
PCBs
500
400
400
300
300
200
200
100
100
0
1960
1970
1980
1990
2000
0
2010
0
1960
1970
1980
1990
2000
0
2010
40
800
DDT
DDT
DDT emissions (kT)
600
30
400
20
200
10
0
1960
1970
1980
1990
2000
Emissions (T or kT)
100
0
2010
P
P
Fig. 3. Temporal trends of a- and b-HCH, 10PCBs and DDT in ringed seal blubber from Lancaster Sound in the Canadian Arctic archipelago. Details on the
P
analysis of HCH, PCBs and DDT compounds can be found in Muir et al. (2000). Emissions for HCH isomers are from Li et al. (2000, 2003), DDT (Li and Li,
2004) and PCBs (Breivik et al., 2002a,b).
higher Kaw values, the DDTs and PCBs are more hydrophobic,
and are removed from the water column on sinking particles as
well as by volatilization. The Arctic Ocean water reservoir
thus responds very quickly to emission changes for these compounds relative to that for the ‘‘swimmers’’.
4. Global fate of POPs
4.1. Biogeochemical cycles and geophysical drivers of
the sinks and reservoirs of POPs
The environmental fate and impact of POPs result from the
interplay of numerous processes including physical transport,
multimedia partitioning and biogeochemical cycles. The environmental partitioning processes of POPs have arguably been
studied the most extensively, from the application of fugacitybased models (Mackay, 1979) to the more recent application
of multiparametric approaches (Goss and Schwarzenbach,
2001). Despite these advances, significant uncertainties remain
with respect to physico-chemical properties, such as HLC that
are needed to quantify POPs’ fluxes in the environment.
The identification of organic carbon (OC) as a key compartment where POPs accumulate (Karickhoff et al., 1979;
Mackay, 1979) was the first recognition of the importance of
biogeochemistry on POPs cycling. For example, OC has
been recognized as the key parameter explaining most of the
spatial, horizontal and vertical, variability for POPs in soils
(Cousins et al., 1999). More recently, and to a lesser degree
soot carbon (SC), has also been identified as a key vector
for the transport and partitioning of POPs in the marine environments (Persson et al., 2002; Lohmann et al., 2005).
However, the importance of carbon dynamics on the fate of
POPs goes beyond merely the fraction of OC (and SC) in sediments and soils. The complexity of the carbon cycle has been
omitted from many POP studies. Oceans and deep lakes provide examples were the carbon cycle plays a key role in the
fate of POPs (Fig. 4) (Larsson et al., 1998; Dachs et al.,
2000; Meijer et al., 2006; Jurado et al., 2007).
Coastal areas are highly relevant in terms of POP cycling
since they are highly populated and at the interface between
open oceans and continents. It has been suggested that continental shelves are important global sinks of PCBs (Jonsson
et al., 2003). However, it is not clear to what degree the storage
of PCBs in coastal sediments is a permanent sink. Sediment
resuspension has been identified as a key process capable of
reintroducing POPs to the water column (Jurado et al., 2007)
and this process could thus prevent POPs deposited to continental shelves from being considered permanent sinks. This
process can be relevant in shallow coastal environments, but
its relevance is not clear on global scales, due to high uncertainties on the remobilization potential of POPs accumulated
in continental shelves at 100 or 200 m depth.
On the global scale, the role of OC cycling can be elucidated in terms of potential reservoirs and sinks (Jurado
et al., 2004; Dalla Valle et al., 2005). Reservoirs are highly dependent on partitioning processes, whereas environmental
sinks largely depend on removal processes of POPs from the
soils, atmosphere and surface oceans. These removal processes can be driven by biogeochemical and/or physical factors. Marine settling fluxes and biodegradation are examples
of POPs’ sinks driven by phytoplankton dynamics and the bacterial loop, respectively. Conversely, advection by oceanic currents, including formation of deep oceanic waters and
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Fig. 4. Schematics of processes affecting POP cycling in the water column and soils which have been studied to some degree (figure adapted from Jurado et al.,
2007 (water) and Cousins et al., 1999 (soil), with permission by Elsevier).
upwelling are examples of physical transport processes with
important implications in POP redistribution and sinks, especially for swimmers (Lohmann et al., 2006b; Wania, 2007).
However, for most legacy POPs, atmospheric transport was
quickly identified as the major mode of long-range transport
and global dispersal. The atmospheric residence time and potential for long-range atmospheric transport has been the focus of
a number of studies (Stroebe et al., 2004, 2006; Fenner et al.,
2005) and is dependent on the gas particle partitioning of the
chemical (Bidleman, 1988). On global scales this will depend
on the geochemistry of the aerosols and the compounds’ physical-chemical properties (Gotz et al., 2007). Furthermore, it has
become clear that there is a close relationship between the water
column’s biological pump and the atmospheric long range transport potential (Dachs et al., 2002; Scheringer et al., 2004).
Several of the emerging POPs are both swimmers and persistent, making them ideal tracers for currents and water mass
movements. Even PCBs have been postulated to serve as good
tracers for deep water formation (Lohmann et al., 2006b), while
they are notoriously difficult to measure (e.g., Lohmann et al.,
2004). In contrast, HCHs are more abundant, but are degraded
more efficiently, too (Harner et al., 1999, 2000). PFAs can be regarded as a newly released batch of time tracers for the transport
efficiency to the Arctic. As the properties of POPs become more
diverse, a truly global understanding of their fate will also require a better grasp of the different biogeochemical cycles and
geophysical processes which will combine to determine their
transport and fate. Furthermore, as the universe of chemicals diversify (polar vs. non-polar, HMW vs. LMW, etc.) the need for
partitioning models accounting for this variability in physical
chemical properties will increase (Goss and Schwarzenbach,
2001; Gotz et al., 2007).
4.2. Are there feedbacks between continental and
oceanic POPs?
Most models focusing on the global cycling of POPs have
both terrestrial and marine environments. However, usually
terrestrial environments are considered the source and reservoir of POPs while oceans and the atmosphere are merely
sinks and transport vectors. The identification of the spatial
and seasonal global potential reservoirs of POPs in oceans,
soils and vegetations has been attempted by Jurado et al.
(2004) and Dalla Valle et al. (2004, 2005). Briefly, these authors suggested that both the spatial distribution and seasonal
cycling of marine and terrestrial organic matter (or soot carbon) play a key role in the spatial distribution of POPs, parallel
to physical controls such that exerted by temperature and its
variability. This is consistent with a number of field studies
(Meijer et al., 2003; Ribes et al., 2003) showing that temperature and OC are drivers of the observed variability of
POPs’ distributions.
Comparing the potential reservoirs of POPs with their
actual measured distribution is very intriguing (Fig. 5). Predictions show that soil maximum reservoir capacities range from
60 N to 80 N, and are mainly a function of OC content and
temperature. Conversely, actual measurements show
maximum soil concentrations at 60 N, ranging from 50 N
to 70 N, which is slightly north of the POPs’ historical usage
(Meijer et al., 2003). Therefore, there are latitudinal differences between sources, measured concentrations and calculated potential reservoirs. Numerous processes could prevent
POPs from moving northwards to reach a distribution mimicking that of the potential reservoirs. These processes potentially
include the retarded grasshopping of POPs over terrestrial environments due to strong sorption (Meijer et al., 2003), atmospheric and oceanic circulation patterns, oceanic and terrestrial
removal processes, and global dynamics that act at the continent-ocean scale.
Is it mere coincidence that the maximum measured concentrations of PCBs in soils are in the same latitudinal range as
the maximum potential reservoir for the oceans, which are, indeed, around 60 N (Fig. 5)? This oceanic capacity is higher during high productivity seasons (spring) and lower during winter.
Water temperatures are never as low as terrestrial ones, and
therefore, oceans could become a source of semi-volatile
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copyright by the American Chemical Society). Center panel shows predicted maximum reservoir capacity of POPs in top soils and surface mixed layer in comparison to that in the atmosphere (from Dalla Valle et al., 2005 with copyright by Elsevier). Right panel shows latitudinal profiles of maximum reservoir capacity
for the Atlantic Ocean (from Jurado et al., 2004 with copyright by Elsevier).
POPs to adjacent continents. The predominant atmospheric circulation patterns move air masses along east-west axes. If this
atmospheric coupling between marine and land storage of
POPs was effective, then the measured latitudinal distribution
of POPs in continental soils mimicking the prediction for the
ocean would be explained. Interactions and feedbacks between
oceans and continents are known to occur for gases such as CO2.
A realistic assessment of these feedbacks for the fate of POPs is
needed. These interactions will undoubtedly be different for
swimmers, flyers and hoppers.
The comparison of POP cycling suggests similar depositional and removal processes for surface ocean and land
(Fig. 4). However, this apparent similarity is fictitious. Significant differences occur between land and ocean. Soil OC has
a much higher capacity than the oceanic mixed layer OC to retain POPs (except in continental deserts), and leaching of OC
and associated POPs is not an effective removal process from
soils either. Conversely, removal processes of POPs from surface oceanic waters are the most effective sink of marine
POPs. Indeed, there is a reasonable number of measurements
of hydrophobic POPs sinking fluxes due to their association
with settling particles by using sediment traps deployments
or Th-U disequilibrium estimation techniques (Tanabe and
Tatsukawa, 1983; Fowler and Knauer, 1986; Dachs et al.,
1996; Gustafsson et al., 1997a; Dachs et al., 1999; Bouloubassi et al., 2006). However, the significance of this process
on the global dynamics of POPs has only lately been assessed
by Dachs et al. (2002). They demonstrated the importance of
the biological pump as an efficient removal process in eutrophic regions, especially for the more hydrophobic POPs.
Once POPs are removed from the open ocean mixed layer,
they are no longer available for global cycling. This removal
process can be more important than other sinks such as atmospheric OH radical degradation reactions (Wania and Daly,
2002).
Few studies have compared the fate of POPs in soils and
sediments simultaneously. High mountain lake sediments
and soils provide good examples, as these two matrices receive
the same atmospheric inputs, but process POPs differently.
PCBs and PAHs display a more degraded signal in the sediments and some evidence for diagenetic PAHs in soils (Grimalt et al., 2004a,b). These studies have shown that in
addition to deposition processes, for which temperature plays
a role, biogeochemical cycling in the water column, soils and
sediments is also important. In addition, the studies have also
highlighted the important role that high mountain regions play
as cold traps of POPs (Grimalt et al., 2001; Daly and Wania,
2005). Recently, Daly and Wania (2005) summarized the different geophysical processes that lead to a convergence of
POPs in high mountain regions, thus enhancing their bioconcentrations in mountain food webs (Vives et al., 2004). In
this respect, more research is needed on the accumulation processes of POPs in these food webs, their biogeochemical controls and how important high mountain regions are as traps for
POPs on the global scale and for chemicals with varying physical chemical properties.
4.3. Global change and POPs
The multiple and varied manifestations of global change
processes that affect the Earth will undoubtedly influence
the fate and potentially the impact of POPs. It is a different
question whether our current knowledge of POP dynamics
and global change scenarios is sufficient to predict the ‘‘global
change of POPs’’. Manifestations of global change include,
but are not limited to, changes in temperature and its variability, losses of biodiversity, changes in the hydrological cycles
and desertification, increase of extreme meteorological events
and social/demographic effects due to the above processes in
a globalized economy. Macdonald et al. (2003) have reviewed
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eloquently some of the variables that can drive a change in
POPs’ fate and impact at the global scale and for the Arctic
Ocean in particular (Macdonald et al., 2005), followed by particular case studies (Ma et al., 2004a,b; Dalla Valle et al.,
2007).
Issues such as changes in the sorbing phases, migration
habits, food web structure or meteorological variables (precipitation patterns, wind speed, temperature, circulation patterns)
will modify POP persistence, long range atmospheric transport, emissions and sinks, degradation processes, bioaccumulation and exposure routes and toxicology. The lack of
appropriate time series measurements has meant that only in
very few cases have changes in POPs dynamics been observed
(e.g., Ma et al., 2004a,b) An important question concerns the
extent to which changes in the biogeochemical OC cycle and
the geophysical drivers will remobilize historical reservoirs of
POPs. In this respect it is important to assess whether flooding,
drought events, and other extreme meteorological events could
modify POPs’ persistence and long range transport potential.
5. Towards global mass balances of POPs
Ideally, we would like to close the mass balance for each
group of POPs, accurately knowing its emissions and accounting for its environmental reservoirs and sink fluxes. However,
as mentioned above, even estimating the emissions of industrially produced POPs remains a challenge. Alternatively, we
could constrain the emissions by properly accounting for the
amount of POPs residing in environmental reservoirs (organic
carbon, water and air) and estimating the sink fluxes. For most
legacy POPs, the carbon cycle presents a unique opportunity to
derive the strength of most environmental reservoirs (see
above). However, accounting for POPs’ environmental sinks
poses its very own challenges. By definition, sinks are final removal fluxes of POPs out of the active cycling reservoirs,
hence they combine degradation reactions in the atmosphere
(photoloysis, radical reactions), transformations by microbes
in soil, sediment and water and chemical reactions in soils/sediments. In addition, the fluxes moving POPs into the deep
ocean, and the deep soil and sediment horizons can constitute
final sinks, as long as no overturning recycles the accumulated
POPs back into the actively cycling environmental reservoirs.
5.1. How global are POPs?
Initially, it seems appropriate to define what ‘global fate’
actually means. For most legacy POPs, previous studies have
highlighted the de facto separation between the Northern
(NH) and Southern Hemisphere (SH) (e.g., Ballschmiter and
Wittlinger, 1991; Ballschmiter, 1992). The NH encompasses
most industrialized nations, and accounts for the majority of
the population on Earth. It is hence not surprising that legacy
POPs, such as PCBs, HCBs and PCDD/Fs display much
higher emissions and concentrations in the NH (e.g., Breivik
et al., 2002a,b; Meijer et al., 2003; Jonsson et al., 2003; Barber
et al., 2005). The atmospheric exchange between the two
hemispheres is slow (on the order of one year), thus a steep
gradient with higher concentrations in the NH and lower
ones in the SH is to be expected.
For examples, most PCBs were used in the NH, and are residing in the NH’s soils and sediments (see also Fig. 5). Surveys into the global distribution of PCDD/Fs in soils again
highlight their preponderance in the NH. For both PCBs and
PCDD/Fs, recent NortheSouth Atlantic cruises do confirm
a profile with higher concentrations in the NH than in the
SH (e.g., Lohmann et al., 2001; Jaward et al., 2004a,b).
However, Jaward et al. (2004a,b) observed unexpectedly
high atmospheric concentrations of PAHs and PCBs south of
the equator, close to Africa. At this time, it is unclear whether
this was due to local/regional sources, or long-range transport.
However, major shifts are underway that will result in
a wider global dispersal of POPs in the future. First, industrial
activities are increasingly moving to Asia, resulting in major
emissions from countries like China and India. Due to the
shifting nature of the monsoons, certain emissions will be
transported into the SH by the monsoons (Semeena et al.,
2006; Dachs et al., 1999). Second, the importance of POPs
emissions that are linked to low-temperature combustion processes (e.g., PAHs, PCNs, possibly PCDD/Fs) will increase for
Asia, Africa and S. America, where most of biomass burning
and forest fires occur (Crutzen and Andrea, 1990; Kasischke
and Penner, 2004). Third, many currently used pesticides are
increasingly used globally, reflecting their major use in agricultural areas in S. America, Africa and Asia (e.g., Pozo
et al., 2006).
Finally, DDT is certainly the biggest exception to the legacy POPs. While DDT is included in the POPs convention,
its use is on-going in many tropical countries to control vector-borne diseases. This will certainly result in DDT being
the first POP for which the dominant primary emissions are
in the tropics, while secondary sources dominate in temperate
climates.
5.2. Global transport of POPs
The occurrence of elevated concentrations of POPs in the
Arctic ecosystem, and the subsequent theory of ‘‘cold condensation’’, led to numerous research projects quantifying the
global transport and fate of POPs. The transport question concerned the exact processes by which POPs were distributed
away from sources under the impact of environmental forcings, such as temperature, and interactions with other environmental compartments, notably vegetation, soil and oceans. In
most cases, the atmosphere has been assumed to be the most
efficient long-range transport medium. For atmospheric transport, gas-particle partitioning rules fate. Gaseous compounds
undergo radical reactions, notably with OH radicals, while
compounds associated with particles are less reactive (Bidleman, 1988).
For non-polar SVOCs, a linear free energy relationship between the gas-particle partitioning coefficient, Kp, and either
the sub-cooled liquid vapor pressure or its Koa seems sufficient
to explain the observed distributions (Harner and Bidleman,
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1998; Pankow, 1998; Gotz et al., 2007). Only the organic
matter content of the aerosol is needed to predict partitioning.
Notable exceptions are PAHs, and possibly PCDD/Fs, which
display enhanced partitioning in the field (Dachs and Eisenreich, 2000; Lohmann and Lammel, 2004). These pyrogenic
compounds are likely adsorbed to black carbon upon their release. As these particles undergo coagulation and reactions,
the adsorbed PAHs seem to be prevented from freely exchanging with the surrounding atmosphere (Lohmann and Lammel,
2004).
For polar compounds, the use of poly-parameter linear free
energy relationships has been advocated. These relationships
are suitable for any compound, as long as the molecular descriptors for both the compounds and the aerosols are known
(Gotz et al., 2007).
5.3. Multi-media global mass balance
For legacy POPs such as HCB, PCBs and PCDD/Fs, global
soils were easily identified as the main environmental reservoir (Brzuzy and Hites, 1996; Meijer et al., 2003; Barber
et al., 2005). A preliminary purportedly global mass balance
for dioxins was based solely on their fluxes to terrestrial soils
(Brzuzy and Hites, 1996). A more complete assessment of
global sinks was undertaken for PCBs. In the case of PCBs
and HCBs, roughly equal amounts reside in the OC fractions
in soils and in shelf sediments (Jonsson et al., 2003; Meijer
et al., 2003; Barber et al., 2005). However, at least in impacted
sediments, black carbon has been found to be a major adsorbent for hydrophobic aromatic and planar compounds (e.g.,
Gustafsson et al., 1997b; Cornelissen et al., 2005; Lohmann
et al., 2005). Nonetheless, the accumulation of PCBs in remote
shelf sediments suggests an efficient delivery of these hydrophobic compounds, presumably via atmospheric transport,
airewater exchange, uptake by particulates and subsequent
settling to the sediments. The importance of the marine sink
has been highlighted by numerous contributions, linking
POPs to the biological pump (Dachs et al., 2000, 2002; Jurado
et al., 2005). For the deeper ocean waters, no inventory has yet
been published, although the presence of dissolved POPs has
been measured for quite some time (Schulz-Bull et al.,
1998). The first truly global measurements were actually performed in water (Iwata et al., 1993). In the remote oceans,
most particulate material settling out of the surface layer
does not reach the ocean floor. This implies that any POPs associated with the particulates at the surface are either degraded
by bacteria or are released at depth. The fate of these settling
POPs in the mesopelagic ocean, with their fate linked to the
important mineralization processes of OC during downward
transport needs further research. For the less persistent
HCHs, there is ample evidence of a selective enantiomeric
degradation in the water column (Harner et al., 1999, 2000).
Debates are still focussing on the importance of the atmospheric sink, i.e., the atmospheric degradation of POPs due
to OH-radical attack as discussed elsewhere (Axelman and
Gustafsson, 2002; Lohmann et al., 2006a). Estimations of
the amount of POPs lost due to atmospheric reactions need
159
to take into account the fraction present in the gas phase, the
temperature-dependency of the reaction and the seasonal and
altitudinal variations in OH-radicals (e.g., Lohmann et al.,
2006a). This is an area where general circulation models might
give best results since important uncertainty exists in the
knowledge for variability of POP concentrations geographically and in the air column. For diurnal studies, the interplay
of OH-radicals, temperature-driven volatilization and the fluctuating mixing height of the atmosphere also need to be considered (MacLeod et al., 2007).
Mass balance approaches seem to suggest that emission of
PCBs may be underestimated (e.g., Breivik et al., 2002a,b).
For PCDD/Fs, a truly global estimation of sinks was published, accounting for PCDD/Fs in soils, atmosphere and the
marine sink. Soils accounted for roughly half of the global
sinks. It appeared that the sink fluxes outweighed the known
source terms of PCDD/Fs (Lohmann et al., 2006a). An interesting conclusion regarding PCDD/Fs might be that the
official source strength estimates provided by individual countries underestimate on-going emissions (e.g., Fiedler, 2007).
Despite best efforts from the UNEP, some sources are difficult
to identify and quantify (Minh et al., 2003; Leung et al., 2007).
There might also be a tendency from the official agencies in
charge of estimating their countries emissions to appear
‘clean’. Possibly more important, emission factors (EFs) are
routinely determined under standard operating conditions,
thereby not reflecting the full duty cycle of a given installation
(from starting up to shutting down). In some cases, PCDD/F
emissions from incinerators were likely underestimated by
a factor of two (Wang et al., 2007). Lastly, for many installations, generic EFs have to been used in the absence of measuring individual installations. It only takes a few installations
with very high real PCDD/F emissions to exceed the calculated representative ones.
Despite many field and modeling studies, major uncertainties remain in closing mass balances for legacy POPs.
This highlights the complexity of the task and indicates that
for emerging POPs such exercises are not yet feasible. Nevertheless, the knowledge gained for legacy POPs gives important
clues on their environmental behavior. Especially the physicalchemical properties were identified that allow a chemical to be
bioaccumulative, persistent and have the potential for long
range transport (see Fig. 1 for an example).
6. Emerging POPs
6.1. Emission pathways
Existing emission inventories have typically emphasized
quantifying atmospheric releases of POPs, because the atmosphere has proven an important and efficient medium at transporting many such substances to remote areas of the globe,
such as the Arctic (e.g., Bidleman et al., 1989; de Wit et al.,
2004). However, a number of emerging POPs may not fall
into the same multimedia region of the chemical space plot
as many legacy POPs (Fig. 1). In such cases, other fate processes may become decisive in controlling their environmental
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fate (Meyer et al., 2005). Furthermore, a better understanding
of the amounts of chemicals that are emitted into environmental media other than air is needed.
As mentioned previously, most emerging POPs tend to be
intentionally produced chemicals, many of which have the potential to be released into the environment from the production, use and disposal of various consumer products, such as
PBDEs (Prevedouros et al., 2004) and PFCAs (Prevedouros
et al., 2006). While the overall human exposure of most legacy
POPs has been assumed to be controlled by environmental exposure through our diet (and hence the focus so far have been
on quantifying environmental releases), recent research in
North America indicates that human exposure in urban areas
from indoor air and dust could be of significance when it
comes to emerging chemicals like PBDEs (e.g., Wilford
et al., 2004) and PFASs (Shoeib et al., 2005). These and related studies have also shown that concentrations of PBDEs,
PFASs and even legacy POPs like PCBs (Currado and Harrad,
1998; Harrad et al., 2006) may be significantly elevated in indoor environments as compared to outdoor environments
(Kohler et al., 2005). Hence, the indoor environment may be
a key release pathway of ‘‘domestic POPs’’ into the outdoor
environment (Currado and Harrad, 1998; Wilford et al.,
2004; Shoeib et al., 2005). Harrad and Diamond (2006) recently hypothesized that such domestic emerging POPs may
over time gradually ‘‘bleed’’ into the environment. Human exposure patterns could shift over time from indoor air and dust
towards exposure through food-chains through a gradual release into our global environment. This would also have important implications for the development of emission
inventories for such chemicals and their changing release pathways over time. There is an urgent need to better understand
the emissions during their entire product lifecycle, including
the compounds’ potential for environmental release from production and manufacturing, indoor as well as outdoor use and
disposal for these emerging chemicals. Hence, a dynamic mass
balance (or ‘‘cradle-to-emissions’’) approach may be the only
feasible approach for determining global emissions of such
substances.
6.2. Fate of emerging POPs
An interesting aspect to the pathways discussion is the fate
of polar degradation products of PFAs, such as perfluorooctane
sulfonates (PFOS) and perfluorooctanoic acid (PFOA) (e.g.
Martin et al., 2004). PFOS and PFOA are anions at environmentally relevant pHs and therefore not readily amenable to
classification by the log Kawelog Koa paradigm. Nevertheless
they are clearly in the ‘‘swimmers’’ category. There are similar
considerations for other polar degradation products such as hydroxy-PCBs (OH-PCBs), which have recently been detected in
precipitation and surface waters (Ueno et al., 2007).
In the case of PFOS and PFOA, it was suggested that their
occurrence in remote Arctic regions can be explained, at least
partially, by the long-range atmospheric transport and oxidation of precursors, such as fluorotelomer alcohols (FTOHs)
and PFASs (e.g., Ellis et al., 2003, 2004; Simcik, 2005; Shoeib
et al., 2006; Wallington et al., 2006). Recent studies have attempted to explain the occurrence of PFOA in the Arctic environment by oceanic transport as a result of manufacture and
use of PFOA in the past (Armitage et al., 2006; Prevedouros
et al., 2006; Wania, 2007). Armitage et al. (2006) assumed
emissions via water, i.e. waste water treatment plants. They
predicted that PFOA concentrations in the Northern Polar
zone (equivalent to the Arctic Ocean) would increase until
about 2030 and then gradually decline as ocean concentrations
adjust to lower emission rates. This scenario is much like that
for b-HCH and reflects redistribution from temperate zones
combined with reduced emissions. Unfortunately temporal
trends of PFOA in Arctic biota are not available because it
does not biomagnify in the marine food chain. Recent measurements of other perfluorocarboxylates (perfluorononanoic,
perfluorodecanoic and perfluoroundecanoic acids), which
have similar sources as PFOA, show increasing concentrations
in ringed seals in Canada (Butt et al., 2007) and Greenland
(Bossi et al., 2005), which are consistent with the modeled
predictions for the 1990s using GloboPOP (Armitage et al.,
2006). However, recent trends in Canadian Arctic ringed seals
for PFOS (which was phased out by the manufacturer in 2001)
show a rapid decline in the past 5 years (Butt et al., 2007). Butt
et al. (2007) have proposed that this rapid decline is due to the
reduction of volatile precursors of PFOS (perfluorosulfonamido alcohols) (Fig. 1B) which are degraded to PFOS in the
atmosphere and enter the marine environment during spring
melt. This pathway, which also occurs for ‘‘multihopping’’
persistent and bioaccumulative chemicals, may be particularly
important for PFAs because they are non-volatile and not revolatilized into the atmosphere during melting, unlike neutral
semivolatile POPs (Wania et al., 1998).
7. Conclusions
Global control strategies aimed at banning production and
new use of intentionally produced POPs may not necessarily
lead to a sharp reduction in emissions because of the potential
long life-time of products containing such persistent chemicals, as exemplified by PCBs (Breivik et al., 2002a,b), or the
continuing use of, e.g. stockpiles of OCPs. The more diverse
the lifecycle and usage pattern of emerging POPs are, the
more challenging it will be to accurately predict their emissions necessary for global mass balances. This will ultimately
require that further developments of emission inventories for
legacy and emerging POPs become a major focus on the research agenda. Emission inventories are an important prerequisite for the development of sound control strategies and
are possibly the key limiting factor in understanding the environmental behavior of such compounds on a global scale. It is
unclear in this context whether CLRTAP, which was originally
developed to mitigate air pollutants, is a suitable framework
for controlling some emerging POPs, if sources and fate
may prove to be intimately linked to the hydrosphere. A similar convention covering the ‘‘swimmers’’ might be needed,
unless the CLRTAP can be modified to encompass all POPs.
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Even though major efforts have been made to study the
global reservoirs and sinks of POPs, especially of PCBs, our
understanding of several key processes is largely incomplete.
For example, regarding the cycling of POPs in the ocean
(Fig. 4), we do not understand how zooplankton and other organisms, which are active swimmers, affect the movement of
POPs in the water column, especially in delivering POPs from
the surface to deep waters. Neither do we know much about
the role of bacteria, and the microbial loop in general, on
POPs’ cycling and degradation, in the oceans, sediments and
soils.
In most current models, the physical characterization of
oceanic vertical structure and transport is overly simplistic.
Recently, Jurado et al. (2007) have shown that the assumption
of well-mixed surface waters is not always fulfilled, with implications on POPs’ fate and transport field studies and modeling. For example, it is not clear that dissolved phase
concentrations measured at 5e10 m depth are useful for
airewater exchange estimations. In terms of modeling, the assumption of well mixed compartments can bias flux estimations and response times. Furthermore, many field studies in
ecosystems with difficult accessibility, such as oceans, do
take place during the summer, with limited coverage of other
seasons. A comprehensive assessment of oceanic circulation,
the role of oceanic eddies, filaments, frontal zones and many
other geophysical structures is also needed.
In addition, there is an urgent need to better constrain the
very persistency that put POPs on the international political
agenda. At current, we have problems accurately quantifying
the various degradation processes that we know occur, such
as photolysis in atmosphere and surface waters, degradation
by bacteria and other biological degradation of POPs during
the carbon cycle. In fact, accounting for the amount of POPs
depleted by OH-radicals is still controversial (e.g., Axelman
and Gustafsson, 2002; Lohmann et al., 2006a), although at
least we know the important reactants, their concentrations
and rate constants.
Furthermore, due to the affinity of POPs to OC, the metabolic status of ecosystems should affect the sources and sinks
of POPs. For example, autotrophic ecosystems (where primary
productivity is higher than respiration) do increase the amount
of organic carbon present, and may represent net sinks of
POPs. Conversely, heterotrophic ecosystems (which are net
sources of CO2 to the atmosphere) could be net secondary
sources of POPs to the atmosphere, thus contributing to their
long-range redistribution. Hence, studies relating POPs cycling to ecosystem metabolism are lacking but could be enlightening in further improving our understanding of how
the carbon cycle interacts with POPs.
Hence, much uncertainty is still present in the estimations
of global sinks and reservoirs of POPs due to imperfect parameterizations, lack of knowledge of potential processes, and incomplete global and seasonal coverage of POP concentrations.
Furthermore, the environmental cycling of ionic or polar compounds has not been as studied as that of hydrophobic compounds, and huge uncertainties exist on their environmental
fate on a global scale.
161
Finally, beyond the research areas outlined above, we still
see a major need for monitoring of legacy and emerging
POPs, in water, air and biota. As pointed out earlier, monitoring has helped to detect trends of POPs and improve our understanding of their sources, occurrence and fate. This will
remain important unless we gain a much improved understanding of source releases. This will become even more important as the group of POPs is growing.
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