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DDT and HCH, two discontinued organochlorine insecticides in the Great Lakes region: Isomer trends and sources
Environment International 69 (2014) 159–165
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DDT and HCH, two discontinued organochlorine insecticides in the Great Lakes region: Isomer trends and sources Marta Venier, Ronald A. Hites ⁎ School of Public and Environmental Affairs, Indiana University, Bloomington, IN 47405, United States
a r t i c l e
i n f o
Article history: Received 23 October 2013 Accepted 27 March 2014 Available online 22 May 2014 Keywords: Dicofol Lindane Atmosphere Insecticides Isomer-specific
a b s t r a c t The uses of the insecticides 1,1′-(2,2,2-trichloroethylidene)bis[4-chlorobenzene] (p,p′-DDT) and γhexachlorocyclohexane (γ-HCH) have been discontinued for several years, but they are still consistently detected in air samples collected on the shores of the Great Lakes. Although the agricultural uses of DDT have been restricted in the United States since 1972, DDT continued to be used to manufacture the miticide, dicofol, up until 2011. The use of the technical HCH mixture in North America was restricted in the 1970s, when it was replaced by one of its purified conformers, γ-HCH, also known as lindane. In this study, we have focused on isomer-specific data to gain insights on the temporal trends and possible sources of these compounds. In particular, we calculated ratios of the concentrations of p,p′-DDE + p,p′-DDD versus the sum of the concentrations of the three p,p′ isomers. These ratios are about the same at all five of our sampling sites and are about the same as observed globally. We also calculated the ratio of the concentrations of o,p′-DDT versus the sum of concentrations of o,p′-DDT + p,p′-DDT. This ratio has increased significantly at all five sites over the last 15–20 years. We suggest that dicofol, which contained about 11% o,p′-DDT, may now be a significant, additional source of DDT to the Great Lakes. The average ratio of the concentration of γ-HCH (lindane) versus the sum of the concentrations of γ-HCH + α-HCH did not vary significantly with time, but it did show an urban signature, suggesting that cities may be more important sources of these compounds than previously suspected. © 2014 Elsevier Ltd. All rights reserved.
1. Introduction DDT has a long history as an environmental contaminant. Technical DDT was first synthetized in 1874, but its insecticidal activity was not discovered until 1939. DDT came to the public's attention just after the Second World War as millions of refugees and former prisonersof-war were displaced throughout Europe and Asia. DDT proved to be effective in killing disease-carrying insects that were sometimes prevalent among these populations. DDT was released for widespread agricultural use in the United States in 1946, and its production peaked in 1963, reaching about 80,000 tons per year in the U.S. (Environmental Health Criteria, 1979). Domestically, DDT was used to kill mosquitoes on public beaches, to kill boll weevils on cotton, to kill caterpillars on urban trees, and to kill many other noxious insects. DDT was almost too good to be true, and indeed, by 1965, it turned out that the benefits of DDT came with an environmental price. For example, DDT interfered with calcium metabolism in birds, and this led to thinner eggshells among higher trophic-order birds, such as eagles and other raptors (Agency for Toxic Substances and Disease Registry, 2002). Because of these and other detrimental environmental ⁎ Corresponding author. E-mail address: [email protected] (R.A. Hites).
http://dx.doi.org/10.1016/j.envint.2014.03.028 0160-4120/© 2014 Elsevier Ltd. All rights reserved.
impacts, DDT was banned for agricultural use in the United States in 1972. In 2004, the Stockholm Convention restricted its use to just the control of disease-carrying insects. In the North American Great Lakes, the ban was effective; for example, total DDT levels (the sum of p,p′-DDT, p,p′-DDE and p,p′-DDD concentrations) in lake trout from Lake Michigan decreased by about a factor of 100 over the period 1970–2010 (Carlson et al., 2010). DDT, as it was used in the environment, was not a chemically pure material. Commercially, DDT was synthesized by the condensation of two moles of chlorobenzene with one mole of chloral (Fig. 1). The resulting technical DDT product contained about 75% of the p,p′-isomer, but the o,p′-isomer was also present in significant amounts (about 15%). Other compounds having four chlorine atoms, DDE and DDD, made up the balance. These two compounds were present as primarily the p,p′-isomers, but small amounts of o,p′-DDE and o,p′-DDD were also in the technical DDT product. Thus, when presenting environmental levels of DDT, it has been important to distinguish which isomer or isomers are being measured. In previous work from our laboratory on DDT's concentrations in the Great Lakes' atmosphere, we have usually presented these data as the sum of the concentrations of p,p′-DDT, p,p′DDE, and p,p′-DDD (notated as ∑DDT) (Venier and Hites, 2010b). In this paper, we present isomer-specific concentrations and trends and extended the previously published series.
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Not all U.S. uses of DDT ended with the 1972 restrictions; in fact, DDT continued to be used as a synthetic intermediate for the production of dicofol (Fig. 1). To produce dicofol, technical DDT was first chlorinated to form Cl-DDT. Because of steric hindrance, o,p′-DDT was not appreciably chlorinated by this reaction; thus, Cl-DDT and the resulting dicofol are almost exclusively the p,p′-isomers. After hydrolysis of Cl-DDT to form dicofol, o,p′-DDT, o,p′-DDE, Cl-DDT, and p,p′-DDT remained in the dicofol product as impurities from the starting material (Qiu et al., 2005). Of these impurities, o,p′-DDT was the most abundant, averaging about 11% by weight of dicofol (Qiu et al., 2005). Dicofol was heavily used in the US. The U.S. Geological Survey Pesticide Survey estimated an annual average use of around 400 metric tons between 1992 and 2000, decreasing to 150 metric tons from 2001 to 2009 (Stone, 2013). In 1986, the US Environmental Protection Agency (EPA) temporarily canceled the use of dicofol until the levels of DDT were reduced to b0.1% (Rasenberg and van den Plassche, 2003), but its use was reinstated in 1987 (Hageman et al., 2006). Dicofol was used in the United States until it was voluntarily withdrawn from the market in 2011 (U.S. Environmental Protection Agency, 2011). The issue of DDT contamination in dicofol has been recognized worldwide. From a regulatory standpoint, the implementation of EU Directive 79/117/EEC by the European Commission strictly limited the DDT content in dicofol to b0.1% (Rasenberg and van den Plassche, 2003). In China, standards requiring DDT impurities to be b0.5% of technical dicofol were supposed to be implemented by 2003, but noncompliant formulations were still available in the Chinese market well after that (Qiu et al., 2005). In terms of environmental detection, there are very few measurements of dicofol in the air. It has been detected in the Arctic (AMAP, 2009; Becker et al., 2012; Zhong et al., 2011) and in Asia (Qiu et al., 2004; Xu et al., 2011; Zhong et al., 2011). There are no measurements of dicofol's atmospheric concentrations in North America. Another insecticide that was widely used on a global scale starting in the 1940s was hexachlorocyclohexane (HCH) (Xiao et al., 2004). This material was a mixture of monocyclic chlorinated hydrocarbons that existed as eight conformers. These were notated as α-HCH through θ-HCH to indicate the sequence in which the axial and equatorial positions on the cyclohexane ring are substituted with chlorine (Willett et al., 1998). The HCH mixture was first discovered in 1825 by Faraday, but its insecticidal properties were not noticed until the 1940s. HCH was
produced commercially by the chlorination of benzene, which produced a mixture of conformers that were collectively known as technical HCH. This material contained about 14% γ-HCH, which is the only conformer with insecticidal properties, 65–70% α-HCH, 7–10% β-HCH, and about 7% δ-HCH (Willett et al., 1998). The other four conformers represented b2% of technical HCH. The use of technical HCH was banned in Canada in 1971 and in the United States in 1978. Technical HCH was replaced on the market by purified (N90%) γ-HCH, which was also called lindane. It has been estimated that, in 1990 in the U.S., the annual usage of lindane was ~ 120 tons (Li et al., 1996). All agricultural uses of γ-HCH were restricted in Canada in 2004 and in the United States in 2009. Now lindane can be found only in therapeutic soaps used to treat head lice in children. The global usage of technical HCH was estimated to be 40,000 tons in 1980 and 29,000 tons in 1990, and for lindane it was 5900 tons in 1980 and 4000 tons in 1990 (Li et al., 1996). HCH has been detected worldwide in all environmental matrices. In particular, HCH has been found in air samples collected in China (Wu et al., 2010), in the Arctic (Becker et al., 2008), and in remote regions such as the Marshall Islands, the Mauritius and southern Chile (Fiedler et al., 2013), just to cite a few examples. In 2009, HCH, both the α- and γ- isomers, as well as lindane, were included in the Stockholm Convention on Persistent Organic Pollutants list, which has led to their phase out. In this paper, we will analyze the atmospheric concentrations of the DDTs and HCHs at five sites around the Great Lakes with an emphasis on the relative isomeric compositions of both DDT and HCH. The purpose of the paper is to report both on the temporal trends of these compounds' concentrations and on the relative amounts of the different DDT congeners and of the different HCH conformers. Our goal is to use these data to identify sources of these compounds and to determine the temporal strengths of these sources. This information will provide insights into the effectiveness of regulatory efforts. 2. Materials and methods 2.1. Sampling and analytical methodology All of the samples were collected as part of the Integrated Atmospheric Deposition Network (IADN), which has been collecting air samples (both the vapor and particle phases) every 12 days at several
Fig. 1. Structures showing the commercial synthesis of DDT and dicofol and the environmental degradation (weathering) of DDT to DDE and DDD. o,p′-DDT is the major impurity (at ~10%) in dicofol.
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Fig. 2. Locations of the atmospheric sampling sites around the North American Great Lakes. The red shaded areas represent population centers.
sites on the shores of the Great Lakes since 1991. The samples were collected at the following United States IADN sites (Fig. 2): Eagle Harbor, Michigan, a remote site on Lake Superior; Sleeping Bear Dunes, Michigan, a very rural site on Lake Michigan; Chicago, Illinois, a highly urbanized site on Lake Michigan; Cleveland, Ohio, an urban site on Lake Erie; and Sturgeon Point, New York, a rural site on Lake Erie (for details, see Environment Canada, 2010). Measurements of these compounds generally started in 1991–1992 at these sites, but the precise timelines over which the measurements were made are given in Fig. S1. Details of the sample collection, extraction, and analysis procedures can be found elsewhere (Salamova and Bayes, 2013; Carlson et al., 2004), and only a brief description will be presented here. A modified Anderson high-volume air sampler (General Metal Works, model GS2310) was used to collect air samples for 24 h at a flow rate giving a total sample volume of about 820 m3. The air stream was first passed through a Whatman quartz fiber filter (QM-A, 20.3 × 25.4 cm) to collect the particles and then through Amberlite XAD-2 resin (Supelco, 20–60 mesh), held in a stainless steel cartridge, to collect the vapor phase components. After sampling, the filters and the XAD were separately Soxhlet extracted for 24 h with 50% (v/v) acetone in hexane. Prior to extraction, the samples were spiked with recovery standards (dibutylchlorendate, δ-HCH, and ε-HCH). The extract was reduced in volume by rotary evaporation, the solvent was exchanged to hexane, and this solution was fractionated on a column containing 3.5% (w/w) water deactivated silica gel. The column was eluted with 25 mL of hexane (fraction 1) and 25 mL of 50% (v/v) hexane in dichloromethane (fraction 2). After N2 blow down, the samples were spiked with internal standards (PCB-65 and PCB-165) that were then used for quantification of the target compounds. The DDTs and HCHs were analyzed by gas chromatography on Agilent Technologies (Palo Alto, CA) 5890 and 6890 instruments equipped with 63Ni electron capture detectors and DB-5 and DB-1701 (J & W Scientific) 60-m columns (250-μm i.d. and 0.1-μm film thickness). Quantitation used the internal standard method. Recovery standards were used to estimate recoveries of each compound in each sample. Results were not recovery corrected because recoveries were in the range 80–120%.
2.2. Statistical calculations All statistical calculations were done with Minitab 16. We calculated ratios of concentrations (see below), which have a higher precision than the concentrations alone. In general, all of these ratios had relative standard deviations of b 10%, which is at least twice as precise as the individual concentration measurements. 2.3. Quality control and quality assurance Quality control and quality assurance procedures were followed to ensure data accuracy. The detailed procedures are described in the IADN Quality Assurance Program Plan and in the IADN Quality Control Project Plan (Basu and Arnold, 2011). 3. Results and discussion The concentrations of these compounds were measured in both the vapor and particle phases for samples collected over different time periods depending on the site and the analyte (see the various sampling timelines in Fig. S1). Using the vapor phase and particle phase data that overlapped in time (for example, p,p′-DDT and α-HCH at Sturgeon Point and Sleeping Bear Dunes for the years 1997–2002, inclusive), we determined that, in general, the fraction of the DDTs in the particle phase was b 15% and that the fraction of α- and γ-HCHs in the particle phase was b3%. These were small amounts, and in fact, this reasoning earlier in the project lead us to abandon the particle phase measurements of pesticide concentrations in May 2003 at all sites except Chicago and Cleveland. For consistency, this paper deals only with the vapor phase concentrations of these compounds. As an additional simplification, the 30 concentrations measured during one year were averaged over that year for each of the compounds in this study. In this way, seasonal effects, which we have shown can be substantial (Venier and Hites, 2010a), were eliminated. These concentration averages are tabulated in the Supporting data and are shown as semilogarithmic plots as a function of time in Figs. S2–S6.
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Many of these concentrations are a function of both the population density at the sampling site and the year of sampling. For example, the atmospheric concentrations of p,p′-DDT at Chicago are much greater than those at Eagle Harbor, and in both cases, the concentrations decrease significantly with time (see Figs. S2 and S6). To model these relationships, we used the following equation to fit the data 2
ln ðC Þ ¼ a0 þ a1 t þ a2 log ðpopÞ
ð1Þ
where t is the sampling year, pop is the number of people living and working within a 25-km radius of the sampling site, a0 is the intercept that rectifies the units, a1 is a rate constant in years−1, and a2 quantitates the effect of population. The time it takes for an atmospheric concentration to decrease by half is called its halving time and is given (in years) by t 1=2 ¼
− ln ð2Þ a1
ð2Þ
The justification for using this model, specifically the population term, has been presented previously by Venier and Hites (2010a,b). The importance of each of the regression parameters in explaining the variability of the concentrations was estimated by calculating the relative contribution of each term to the total variability. This value is reported as the percentage of the sum of squares (SOS) attributed to each factor relative to the total sum of squares. The sum of squares for each term adds up to the model (or regression) sum of squares, allowing us to express the importance of each term as a percent of the overall r2 value. Table 1 gives a summary of the regression results for all of the compounds discussed here. Table S6 gives the complete regression output. 3.1. DDT and related compounds Table 1 shows that the regressions using Eq. (1) are highly significant for all of the DDT-related compounds with F N 47. Four compounds, p,p′-DDT, p,p′-DDE, o,p′-DDT, and ∑DDT, give similar overall regression results. The r2 values average about 86%, the population terms average about 0.071 and represent about 81% of the total sum-ofsquares, and with the exception of o,p′-DDT, the time terms average about −0.082 year−1 and represent 2–9% of the total sum-of-squares. This time term converts to a halving time of 8–9 years. This concordance suggests that these four compounds are behaving similarly to
one another in the environment and that our previous use of ∑ DDT to characterize this behavior was appropriate. The other two compounds of this group, p,p′-DDD and o,p′-DDD, gave regression results that were similar to each other but different than the other four compounds in this group. The r2 values averaged about 65%, the population terms were also lower, at about 0.034 (representing ~ 16% of the total sum-of-squares), and the time terms were much higher, at about −0.21 year−1 (representing about 49% of the total sum-of-squares). This average time term converts to a halving time of 3–4 years. Why do these two DDD-related compounds behave so much differently in the atmosphere compared to the DDT- and DDE-related compounds? The difference in halving times between DDE and the DDD isomers might be related to different DDT degradation pathways: Under aerobic conditions, DDT is degraded to DDE, the most persistent of all DDT related compounds, but under anaerobic conditions it degrades to DDD. The first pathway is the most prevalent and DDE is often resistant to biodegradation, providing a possible explanation for the longer halving times of DDE versus DDD (Ricking and Schwarzbauer, 2012; Spencer et al., 1996; Strompl and Thiele, 1997; Zitko, 2003). In the regressions for p,p′-DDT, p,p′-DDE, o,p′-DDT, and ∑ DDT, almost all of the variability is accounted for in the population term. Fig. 3 shows this part of the regression as the logarithm of the median concentrations as a function of the population term. These are all significant positive relationships, indicating that the median concentrations of these four DDT-related compounds are much higher in the cities of Chicago and Cleveland than at the rural and remote sites. This is surprising given that DDT had been used as an agricultural insecticide, and there has been little agriculture in the cities. It is likely that this urbaneffect was caused by the use of DDT in cities in the 1950s and 1960s for insect control in shade trees that lined many streets at that time. Apparently, DDT is being released from urban soil [or perhaps from treated buildings as suggested by Braeuner et al. (2011) and Dirtu et al. (2012)] into the Great Lakes' atmosphere. This relationship with population ultimately indicates that if the population would decline, the concentration would also decline; at the moment, though, it is not possible to speculate on how long it might take to see this effect since we do not understand environmental hysteresis. One way of determining the sources and origins of DDT in environmental samples is to calculate isomer ratios among DDT and one or more of its environmental degradation products: DDE and DDD. For example, one could calculate the ratio of p,p′-DDE to p,p′-DDT to get an estimate on how long the DDT has been in the environment. This
Table 1 Regression parameters for annual average concentrations of several DDT and HCH-related compounds and their ratios.
Constant term Standard error Probability Time term Standard error Probability %SOS Halving time, t1/2 Standard error, t1/2 Population term Standard error Probability %SOS r2 Prob, r2 N F
p,p′-DDT
p,p′-DDE
o,p′-DDT
∑DDT
p,p′-DDD
o,p′-DDD
Rp,p′
Ro,p′
α-HCH
β-HCH
γ-HCH
Rγ
175 13 b0.1% −0.0874 0.0066 b0.1% 8.9% 7.9 0.6 0.0652 0.0028 b0.1% 79.6% 88.5% b0.1% 81 300.
149 21 b0.1% −0.0745 0.0107 b0.1% 2.2% 9.3 1.3 0.0770 0.0046 b0.1% 76.8% 79.0% b0.1% 81 146
NS NS NS NS NS NS NS NS NS 0.0733 0.0037 b0.1% 86.6% 88.7% b0.1% 54 199
168 15 b0.1% −0.0832 0.0073 b0.1% 6.7% 8.3 0.7 0.0670 0.0031 b0.1% 79.9% 86.6% b0.1% 81 251
320 29 b0.1% −0.160 0.015 b0.1% 48.8% 4.3 0.4 0.0331 0.0063 b0.1% 13.6% 62.4% b0.1% 80 63.8
500. 54 b0.1% −0.250 0.027 b0.1% 48.7% 2.8 0.3 0.0340 0.0066 b0.1% 18.3% 67.0% b0.1% 50 47.8
NS NS NS NS NS NS NS – – 0.00247 0.00075 0.2% 11.8% 14.9% 0.2% 81 6.83
−37.7 5.0 b0.1% 0.0189 0.0025 b0.1% 46.9% – – 0.00253 0.00062 b0.1% 13.0% 59.9% b0.1% 54 38.1
397 12 b0.1% −0.196 0.006 b0.1% 93.3% 3.5 0.1 NS NS NS NS 93.6% b0.1% 81 569
443 42 b0.1% −0.221 0.021 b0.1% 49.2% 3.1 0.3 0.0393 0.0051 b0.1% 28.3% 77.5% b0.1% 51 82.5
407 17 b0.1% −0.202 0.008 b0.1% 78.9% 3.4 0.1 0.0285 0.0035 b0.1% 9.6% 88.5% b0.1% 81 301
NS NS NS NS NS NS NS – – 0.00645 0.00040 b0.1% 70.8% 78.8% b0.1% 81 145
The ratios Rp,p′, Ro,p′, and Rγ are defined by Eqs. (3), (5), and (7). The regression parameters for the three terms were determined from Eq. (1), (4), (6), or (8) as appropriate. Halving times (t1/2, in years) were calculated using Eq. (2). See the Supporting data for data plots and for detailed regression results. “NS” means not significant at P b 1%.
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Median concentration (pg/m3)
100
ΣDDT (0.809) 10
p,p'-DDE (0.751) 1
p,p'-DDT (0.860) o,p'-DDT (0.974) 0.1 0
10
20
30
40
50
log2 (pop) Fig. 3. Median vapor-phase concentrations of p,p′-DDT, p,p′-DDE, and o,p′-DDT, and ∑DDT as a function of the squared logarithm of the number of people living and working within a 25-km radius of the sampling site. ∑DDT is the sum of the p,p′-DDT, p,p′-DDE, and p,p′-DDD concentrations. The numbers in parentheses are the r2 values for each regression; all are significant at P b 6%.
environmental degradation process is frequently called “weathering.” In this case, we will use the following unitless ratio:
Rp;p′
p; p′‐DDE þ p; p′‐DDD ¼ p; p′‐DDE þ p; p′‐DDD þ p; p′‐DDT
ð3Þ
where [p,p′-DDT], [p,p′-DDD], and [p,p′-DDE] are the concentrations (in pg/m3) of each isomer in the atmospheric vapor phase. This ratio has been normalized, compared to the one commonly used in the literature (Li et al., 2007; Liu et al., 2009; Munoz-Arnanz and Jimenez, 2011; Park et al., 2011; Qiu et al., 2004; Wang et al., 2007; Yang et al., 2008), to set its range between 0 and 1, rather than between 0 and infinity. These ratios were calculated for each site and for each sampling year, and are all given in the Supporting data. To investigate the effect of time and of population density around the sampling site, the following regression was used 2
Rp;p′ ¼ b0 þ b1 t þ b2 log ðpopÞ
ð4Þ
0.8
163
where the bi parameters correspond to the ai parameters in Eq. (1). The regression parameters are given in Table 1. This analysis showed a small, but significant, overall correlation (r2 = 14.9%), and the only significant term was the one associated with population. This part of the regression is plotted in Fig. 4. Given this lack of a strong relationship with time or population, we have simply taken the average of all Rp,p′ values across all five sites, and this average is 0.648 ± 0.011. This is similar to values in the literature for air samples collected near Lake Taihu in China (0.48–0.76) (Qiu et al., 2004) and for air samples collected in remote southwestern China (0.65 ± 0.39) (Xu et al., 2011). Ratios suggestive of dicofol-type DDT sources were reported in the Arctic (Becker et al., 2012; Bidleman et al., 2013). There are no precise guidelines on how to interpret Rp,p′ values, but higher values suggest the presence of significant weathering of DDT. An apparent global convergence of Rp,p′ values at 0.6–0.7 may indicate that a rough equilibrium between the rate of degradation of p,p′-DDT to p,p′-DDE and p,p′-DDD and their subsequent degradation (perhaps to monocyclic compounds) has been reached. It is also possible that there is an environmental source of DDT other than technical DDT, and that source could be dicofol (Qiu et al., 2005). One way to distinguish DDT inputs from dicofol inputs is to assess the concentrations of o,p′-DDT, which is the most abundant impurity in technical grade dicofol, relative to the concentrations of p,p′-DDT itself (Qiu et al., 2005). Thus, we define another ratio to compare the relative abundances of the o,p′- to p,p′-DDT isomers: o; p′‐DDT Ro;p′ ¼ o; p′‐DDT þ p; p′‐DDT
ð5Þ
where o,p′-DDT and p,p′-DDT are the concentration of each isomer in the vapor phase (pg/m3). As before, we investigated the effect of time and the site's population density using the following regression 2
Ro;p′ ¼ c0 þ c1 t þ c2 log ðpopÞ
ð6Þ
where the ci parameters correspond to the ai parameters in Eq. (1). The resulting regression parameters are given in Table 1. In this case, all three of the fitted parameters are significant, and the overall r2 value is 60%, which indicates a strong relationship. Of the total sum-ofsquares, most (47%) is from the time-term, and this relationship is shown specifically in Fig. 5. This observation is different than that observed for Rp,p′, which showed no significant temporal variations. This increase in the relative abundance of the o,p′-DDT with passing year is unusual, and warrants a bit more examination. Can we distinguish the situation in which the absolute concentration of dicofol has
0.6 0.5
Rp,p' (0.069)
0.4
Average Ro,p'
Average Rp,p' or Rγ
0.6
0.2
Rγ (0.981)
0.4 0.3 0.2
Ro,p' (0.699)
0.1
0.0 0
10
20
30
40
50
log2 (pop) Fig. 4. Relationship between the overall average Rp,p′ and Rγ values (see Eqs. (3) and (7) for definitions of these ratios) at each site and the population living and working within 25 km of that site. The standard errors of each average are smaller than the symbols. The numbers in parentheses are the r2 values for each regression. The Rp,p′ regression is not statistically significant, but that for Rγ is significant at P b 0.1%.
0.0 1998
2000
2002
2004
2006
2008
2010
2012
Fig. 5. Annual averages of Ro,p′ (see Eq. (2)) in air sampled near the Great Lakes as a function of sampling year. The averages were over the five sites for each year. The error bars are standard errors. The number in parentheses is the r2 value of the regression, which is significant at P b 0.2%.
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increased as opposed to the situation in which the relative concentration of dicofol has increased? The temporal data for o,p′-DDT shown in Figs. S2–S6 show that there is no relationship between the concentrations of this compound and time, but there is a strong relationship between the concentrations of p,p′-DDT and time. These observations suggest that the changes in Ro,p′ are driven by the concentration of p,p′-DDT in the denominator of Eq. (5) and that the amount of dicofol in the atmosphere (as indirectly measured by o,p′-DDT) has remained relatively constant over time while the amount of p,p′-DDT has decreased more rapidly. Because the vapor pressures of o,p′-DDT and p,p′-DDT are different, it is not possible to interpret the environmental value of Ro,p′ in terms of the relative inputs of DDT and dicofol, but it is clear from the rate at which Ro,p′ has increased (as represented by the positive slope in Fig. 5) that the relative concentration of dicofol in Great Lakes air has about doubled over the last decade. This phenomenon has also been observed for Taihu Lake in China (Qiu et al., 2004). Very little data on air measurements of dicofol exists. It was detected in the Arctic (AMAP, 2009) and in samples collected on a cruise from East Asia to the High Arctic (Zhong et al., 2011), hinting that long range transport might be playing a role in the observed levels. Because dicofol was retired from the United States market in 2011, we expect to see a slow decrease of Ro,p′ back to values closer to 0.2 over the next several decades. In fact, the two most recent data points in Fig. 5, corresponding to years 2009 and 2010, are below the regression line, suggesting that this trend may have already started; of course, more data are needed to confirm this suggestion. 3.2. Hexachlorocyclohexane (HCH) The use of this insecticide has long been eliminated. Nevertheless, HCH isomers are still detected in air samples from around the Great Lakes, but perhaps as a result of the restrictions put on their use, these concentrations are now decreasing rapidly. The resulting regression parameters for the three HCH conformers in our dataset are given in Table 1. In this case, the time terms are significant for all three conformers. The overall r2 values are N77%, which indicates a strong relationship, and most of the total sums-of-squares are from the time term. From this term, halving times for α-, β-, and γ-HCH are calculated to be 3.1–3.5 years. This agrees with results we have reported before for α- and γ-HCH (Venier and Hites, 2010a), but results for β-HCH have not been reported before. There are no significant differences in the halving times among the three compounds. For β- and γ-HCH, the population terms are significant, averaging at 0.034, and represent 10–28% of the total sum-of-squares. Given that both of these terms are positive, it is likely that there are some urban sources of β- and γ-HCH, but not of α-HCH, to the environment. As with DDT, the ratio of the HCH isomers can be used to infer useful information regarding the sources and environmental behavior of these compounds. Since α- and γ-HCH were the major components of the HCH technical mixture, we calculated the unitless fraction of the γ isomer Rγ, which is defined as Rγ ¼
½γ‐HCH ½α‐HCH þ ½γ‐HCH
ð7Þ
where [α-HCH] and [γ-HCH] are the concentrations (pg/m3) measured in the atmospheric vapor phase at each of the IADN sites. Like the previous ratios, we modified the equation from the one that is usually reported in the literature (Qiu et al., 2004; Willett et al., 1998; Xu et al., 2011) to limit its range to between 0 and 1, rather than between 0 and infinity. As before, we investigated the effect of sampling year and the site's population density using the following regression 2
Rγ ¼ d0 þ d1 t þ d2 log ðpopÞ
ð8Þ
where the di parameters correspond to the ai parameters in Eq. (1). The resulting regression parameters are given in the rightmost column of Table 1; the only significant parameter was the one for population. This effect is shown in Fig. 4. Clearly, there is a strong positive relationship between the Rγ value at a given site and the number of people living or working within a 25-km radius of that site. This relationship has not been observed before. We have reported previously that the atmospheric concentrations of neither α-HCH nor γ-HCH were significantly dependent on the local human population (Venier and Hites, 2010b), but the results reported here indicate that Rγ increases as population near the sampling site increases. This seeming conundrum may be the result of comparing absolute concentrations to relative concentrations. These measurements of Rγ are more sensitive to human population effects than the absolute measurements of the concentrations themselves, particularly once the Rγ values are averaged over all years. This finding suggests that technical HCH (high in α-HCH) was used almost exclusively as an agricultural insecticide, but that lindane (high in γ-HCH) may have had some urban uses. The value of Rγ has been reported to be in the range of 0.12 to 0.25 for technical HCH (Willett et al., 1998; Xu et al., 2011), and this is about what we observe at the remote site of Eagle Harbor. 4. Conclusions The concentrations of DDT and related compounds are decreasing in the atmosphere around the North American Great Lakes, but there is some evidence that the DDT found as an impurity in dicofol is now contributing to some of this atmospheric load. There is also some evidence that DDT levels in cities are higher than previously suspected. The atmospheric concentrations of α-, β-, and γ-HCH (lindane) are decreasing rapidly, but there is also evidence that the concentrations of lindane are higher in the cities than previously suspected. The overall implications are that compounds such as DDT and lindane, which were previously thought to have exclusively agricultural uses, may also have been used in cities for insect control and that cities may be significant sources of these compounds to the atmosphere. Acknowledgments We thank the entire Team IADN at Indiana University for sample and data acquisition and the U.S. Environmental Protection Agency's Great Lakes National Program Office for funding (Grant No. GL-00E76601-3, Todd Nettesheim, project officer). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.envint.2014.03.028. References Agency for Toxic Substances and Disease Registry. Toxicological profile for DDT, DDE, and DDE; 2002 [September]. Arctic Monitoring and Assessment Programme (AMAP). Arctic pollution 2009; 2009 [Oslo]. Basu I, Arnold K. Integrated Atmospheric Deposition Network Quality Assurance Project Plan. Bloomington, IN: Indiana University; 2011 (http://www.epa.gov/greatlakes/ monitoring/air2/iadn/qapp/iadn-qapp-201111.pdf). Becker S, Halsall CJ, Tych W, Kallenborn R, Su Y, Hung H. Long-term trends in atmospheric concentrations of alpha- and gamma-HCH in the Arctic provide insight into the effects of legislation and climatic fluctuations on contaminant levels. Atmos Environ 2008;42(35):8225–33. Becker S, Halsall CJ, Tych W, Kallenborn R, Schlabach M, Mano S. Changing sources and environmental factors reduce the rates of decline of organochlorine pesticides in the Arctic atmosphere. Atmos Chem Phys 2012;12:4033–44. Bidleman TF, Kurt-Karakus PB, Wong F, Alegria HA, Jantunen LM, Hung H. Is there still “new” DDT in North America? An Investigation using proportions of DDT compounds. In: McConnell LL, Dachs J, Hapeman CJ, editors. Occurrence, Fate and Impact of Atmospheric Pollutants on Environmental and Human Health, vol. 1149. Washington: Amer Chemical Soc; 2013. p. 153–81.
M. Venier, R.A. Hites / Environment International 69 (2014) 159–165 Braeuner EV, Mayer P, Gunnarsen L, Vorkamp K, Raaschou-Nielsen O. Occurrence of organochlorine pesticides in indoor dust. J Environ Monit 2011;13:522–6. Carlson DL, Basu I, Hites RA. Annual variations of pesticide concentrations in Great Lakes precipitation. Environ Sci Technol 2004;38:5290–6. Carlson DL, Vault DSD, Swackhamer DL. On the rate of decline of persistent organic contaminants in lake trout (Salvelinus namaycush) from the Great Lakes, 1970–2003. Environ Sci Technol 2010;44:2004–10. Dirtu AC, Ali N, Van den Eede N, Neels H, Covaci A. Country specific comparison for profile of chlorinated, brominated and phosphate organic contaminants in indoor dust. Case study for Eastern Romania, 2010. Environ Int 2012;49:1–8. Environment Canada. http://www.ec.gc.ca/rs-mn/default.asp?lang=En&n=BFE9D3A3-1, 2010. [accessed 4 September 2013]. Environmental health criteria 9: DDT and its derivatives. World Health Organization; 1979. Fiedler H, Abad E, van Bavel B, de Boer J, Bogdal C, Malisch R. The need for capacity building and first results for the Stockholm Convention Global Monitoring Plan. Trends Anal Chem 2013;46:72–84. Hageman KJ, Simonich SL, Campbell DH, Wilson GR, Landers DH. Atmospheric deposition of current-use and historic-use pesticides in snow at national parks in the western United States. Environ Sci Tech 2006;40(10):3174–80. Li Y-F, McMillan A, Scholtz MT. Global HCH usage with 1° × 1° longitude/latitude resolution. Environ Sci Technol 1996;30(12):3525–33. Li J, Zhang G, Guo L, Xu W, Li X, Lee CSL, Ding A, Wang T. Organochlorine pesticides in the atmosphere of Guangzhou and Hong Kong: regional sources and long-range atmospheric transport. Atmos Environ 2007;41(18):3889–903. Liu X, Zhang G, Li J, Yu LL, Xu Y, Li XD, Kobara Y, Jones KC. Seasonal patterns and current sources of DDTs, chlordanes, hexachlorobenzene, and endosulfan in the atmosphere of 37 Chinese cities. Environ Sci Tech 2009;43(5):1316–21. Munoz-Arnanz J, Jimenez B. New DDT inputs after 30 years of prohibition in Spain. A case study in agricultural soils from south-western Spain. Environ Pollut 2011;159: 3640–6. Park JS, Shin SK, Kim WI, Kim BH. Residual levels and identify possible sources of organochlorine pesticides in Korea atmosphere. Atmos Environ 2011;45(39):7496–502. Qiu XH, Zhu T, Jing L, Pan HS, Li QL, Miao GF, Gong JC. Organochlorine pesticides in the air around the Taihu Lake, China. Environ Sci Technol 2004;38:1368–74. Qiu XH, Zhu T, Yao B, Hu JX, Hu SW. Contribution of dicofol to the current DDT pollution in China. Environ Sci Technol 2005;39:4385–90. Rasenberg MHC, van den Plassche EJ. Risk profile and summary report for dicofol dossier prepared for the UNECE Convention on Long-range Transboundary Air Pollution's expert group on POPs; 2003 [Reference: 4 L0002.A1/R0012/EVDP/Nijm]. Ricking M, Schwarzbauer J. DDT isomers and metabolites in the environment: an overview. Environ Chem Lett 2012;10(4):317–23.
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Salamova A, Bays JC. Collection of air and precipitation samples: IADN project sample procedure. Bloomington, IN: Indiana University; 2013 (http://www.epa.gov/greatlakes/ monitoring/air2/iadn/qapp/iadn-sop-collection-air-precipitation-samples-201303. pdf). Spencer WF, Singh G, Taylor CD, LeMert RA, Cliath MM, Farmer WJ. DDT persistence and volatility as affected by management practices after 23 years. J Environ Qual 1996;25: 815–21. Stone WW. Estimated annual agricultural pesticide use for counties of the conterminous United States, 1992–2009. U.S. Geological Survey Data Series, 752; 2013 [1-p. pamphlet, 14 tables]. Strompl C, Thiele J. Comparative fate of 1,1-diphenylethylene (DPE), 1,1-dichloro-2,2bis(4chlorophenyl)-ethylene (DDE), and pentachlorophenol (PCP) under alternating aerobic and anaerobic conditions. Arch Environ Contam Toxicol 1997;33:350–6. U.S. Environmental Protection Agency. http://www.epa.gov/opp00001/reregistration/ dicofol/dicofol-moa-5-17-11.pdf, 2011. [accessed 4 September 2013]. Venier M, Hites RA. Time trend analysis of atmospheric POPs concentrations in the Great Lakes region since 1990. Environ Sci Technol 2010a;44:8050–5. Venier M, Hites RA. Regression model of partial pressures of PCBs, PAHs, and organochlorine pesticides in the Great Lakes' atmosphere. Environ Sci Technol 2010b;44: 618–23. Wang J, Guo L, Li J, Zhang G, Lee CSL, Li X, Jones KC, Xiang Y, Zhong L. Passive air sampling of DDT, chlordane and HCB in the Pearl River Delta, South China: implications to regional sources. J Environ Monit 2007;9(6):582–8. Willett KL, Ulrich EM, Hites RA. Differential toxicity and environmental fates of hexachlorocyclohexane isomers. Environ Sci Technol 1998;32:2197–207. Wu X, Lam JCW, Xia C, Kang H, Sun L, Xie Z, Lam PKS. Atmospheric HCH Concentrations over the Marine Boundary Layer from Shanghai, China to the Arctic Ocean: role of human activity and climate change. Environ Sci Technol 2010;44(22):8422–8. Xiao H, Li N, et al. Compilation, evaluation, and selection of physical–chemical property data for α-, β-, and γ-hexachlorocyclohexane. J Chem Eng Data 2004; 49(2):173–85. Xu Y, Zhang G, Li J, Chakraborty P, Li H, Liu X. Long-range atmospheric transport of persistent organochlorinated compounds from south and mainland south-eastern Asia to a remote mountain site in south-western China. J Environ Monit 2011;13: 3119–27. Yang Y, Li D, Mu D. Levels, seasonal variations and sources of organochlorine pesticides in ambient air of Guangzhou, China. Atmos Environ 2008;42(4):677–87. Zhong G, Xie Z, Cai M, Möller A, Sturm R, Tang J, Zhang G, He J, Ebinghaus R. Distribution and air–sea exchange of current-use pesticides from East Asia to the high Arctic Ocean. Environ Sci Technol 2011;46(1):259–67. Zitko V. Chlorinated pesticides: aldrin, DDT, endrin, dieldrin, mirex. In: Fiedler H, editor. Persistent Organic Pollutants, vol. 30. Heidelberg: Springer Berlin; 2003. p. 47–90.
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DDT and HCH, two discontinued organochlorine insecticides in the Great Lakes region: Isomer trends and sources Marta Venier, Ronald A. Hites ⁎ School of Public and Environmental Affairs, Indiana University, Bloomington, IN 47405, United States
a r t i c l e
i n f o
Article history: Received 23 October 2013 Accepted 27 March 2014 Available online 22 May 2014 Keywords: Dicofol Lindane Atmosphere Insecticides Isomer-specific
a b s t r a c t The uses of the insecticides 1,1′-(2,2,2-trichloroethylidene)bis[4-chlorobenzene] (p,p′-DDT) and γhexachlorocyclohexane (γ-HCH) have been discontinued for several years, but they are still consistently detected in air samples collected on the shores of the Great Lakes. Although the agricultural uses of DDT have been restricted in the United States since 1972, DDT continued to be used to manufacture the miticide, dicofol, up until 2011. The use of the technical HCH mixture in North America was restricted in the 1970s, when it was replaced by one of its purified conformers, γ-HCH, also known as lindane. In this study, we have focused on isomer-specific data to gain insights on the temporal trends and possible sources of these compounds. In particular, we calculated ratios of the concentrations of p,p′-DDE + p,p′-DDD versus the sum of the concentrations of the three p,p′ isomers. These ratios are about the same at all five of our sampling sites and are about the same as observed globally. We also calculated the ratio of the concentrations of o,p′-DDT versus the sum of concentrations of o,p′-DDT + p,p′-DDT. This ratio has increased significantly at all five sites over the last 15–20 years. We suggest that dicofol, which contained about 11% o,p′-DDT, may now be a significant, additional source of DDT to the Great Lakes. The average ratio of the concentration of γ-HCH (lindane) versus the sum of the concentrations of γ-HCH + α-HCH did not vary significantly with time, but it did show an urban signature, suggesting that cities may be more important sources of these compounds than previously suspected. © 2014 Elsevier Ltd. All rights reserved.
1. Introduction DDT has a long history as an environmental contaminant. Technical DDT was first synthetized in 1874, but its insecticidal activity was not discovered until 1939. DDT came to the public's attention just after the Second World War as millions of refugees and former prisonersof-war were displaced throughout Europe and Asia. DDT proved to be effective in killing disease-carrying insects that were sometimes prevalent among these populations. DDT was released for widespread agricultural use in the United States in 1946, and its production peaked in 1963, reaching about 80,000 tons per year in the U.S. (Environmental Health Criteria, 1979). Domestically, DDT was used to kill mosquitoes on public beaches, to kill boll weevils on cotton, to kill caterpillars on urban trees, and to kill many other noxious insects. DDT was almost too good to be true, and indeed, by 1965, it turned out that the benefits of DDT came with an environmental price. For example, DDT interfered with calcium metabolism in birds, and this led to thinner eggshells among higher trophic-order birds, such as eagles and other raptors (Agency for Toxic Substances and Disease Registry, 2002). Because of these and other detrimental environmental ⁎ Corresponding author. E-mail address: [email protected] (R.A. Hites).
http://dx.doi.org/10.1016/j.envint.2014.03.028 0160-4120/© 2014 Elsevier Ltd. All rights reserved.
impacts, DDT was banned for agricultural use in the United States in 1972. In 2004, the Stockholm Convention restricted its use to just the control of disease-carrying insects. In the North American Great Lakes, the ban was effective; for example, total DDT levels (the sum of p,p′-DDT, p,p′-DDE and p,p′-DDD concentrations) in lake trout from Lake Michigan decreased by about a factor of 100 over the period 1970–2010 (Carlson et al., 2010). DDT, as it was used in the environment, was not a chemically pure material. Commercially, DDT was synthesized by the condensation of two moles of chlorobenzene with one mole of chloral (Fig. 1). The resulting technical DDT product contained about 75% of the p,p′-isomer, but the o,p′-isomer was also present in significant amounts (about 15%). Other compounds having four chlorine atoms, DDE and DDD, made up the balance. These two compounds were present as primarily the p,p′-isomers, but small amounts of o,p′-DDE and o,p′-DDD were also in the technical DDT product. Thus, when presenting environmental levels of DDT, it has been important to distinguish which isomer or isomers are being measured. In previous work from our laboratory on DDT's concentrations in the Great Lakes' atmosphere, we have usually presented these data as the sum of the concentrations of p,p′-DDT, p,p′DDE, and p,p′-DDD (notated as ∑DDT) (Venier and Hites, 2010b). In this paper, we present isomer-specific concentrations and trends and extended the previously published series.
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Not all U.S. uses of DDT ended with the 1972 restrictions; in fact, DDT continued to be used as a synthetic intermediate for the production of dicofol (Fig. 1). To produce dicofol, technical DDT was first chlorinated to form Cl-DDT. Because of steric hindrance, o,p′-DDT was not appreciably chlorinated by this reaction; thus, Cl-DDT and the resulting dicofol are almost exclusively the p,p′-isomers. After hydrolysis of Cl-DDT to form dicofol, o,p′-DDT, o,p′-DDE, Cl-DDT, and p,p′-DDT remained in the dicofol product as impurities from the starting material (Qiu et al., 2005). Of these impurities, o,p′-DDT was the most abundant, averaging about 11% by weight of dicofol (Qiu et al., 2005). Dicofol was heavily used in the US. The U.S. Geological Survey Pesticide Survey estimated an annual average use of around 400 metric tons between 1992 and 2000, decreasing to 150 metric tons from 2001 to 2009 (Stone, 2013). In 1986, the US Environmental Protection Agency (EPA) temporarily canceled the use of dicofol until the levels of DDT were reduced to b0.1% (Rasenberg and van den Plassche, 2003), but its use was reinstated in 1987 (Hageman et al., 2006). Dicofol was used in the United States until it was voluntarily withdrawn from the market in 2011 (U.S. Environmental Protection Agency, 2011). The issue of DDT contamination in dicofol has been recognized worldwide. From a regulatory standpoint, the implementation of EU Directive 79/117/EEC by the European Commission strictly limited the DDT content in dicofol to b0.1% (Rasenberg and van den Plassche, 2003). In China, standards requiring DDT impurities to be b0.5% of technical dicofol were supposed to be implemented by 2003, but noncompliant formulations were still available in the Chinese market well after that (Qiu et al., 2005). In terms of environmental detection, there are very few measurements of dicofol in the air. It has been detected in the Arctic (AMAP, 2009; Becker et al., 2012; Zhong et al., 2011) and in Asia (Qiu et al., 2004; Xu et al., 2011; Zhong et al., 2011). There are no measurements of dicofol's atmospheric concentrations in North America. Another insecticide that was widely used on a global scale starting in the 1940s was hexachlorocyclohexane (HCH) (Xiao et al., 2004). This material was a mixture of monocyclic chlorinated hydrocarbons that existed as eight conformers. These were notated as α-HCH through θ-HCH to indicate the sequence in which the axial and equatorial positions on the cyclohexane ring are substituted with chlorine (Willett et al., 1998). The HCH mixture was first discovered in 1825 by Faraday, but its insecticidal properties were not noticed until the 1940s. HCH was
produced commercially by the chlorination of benzene, which produced a mixture of conformers that were collectively known as technical HCH. This material contained about 14% γ-HCH, which is the only conformer with insecticidal properties, 65–70% α-HCH, 7–10% β-HCH, and about 7% δ-HCH (Willett et al., 1998). The other four conformers represented b2% of technical HCH. The use of technical HCH was banned in Canada in 1971 and in the United States in 1978. Technical HCH was replaced on the market by purified (N90%) γ-HCH, which was also called lindane. It has been estimated that, in 1990 in the U.S., the annual usage of lindane was ~ 120 tons (Li et al., 1996). All agricultural uses of γ-HCH were restricted in Canada in 2004 and in the United States in 2009. Now lindane can be found only in therapeutic soaps used to treat head lice in children. The global usage of technical HCH was estimated to be 40,000 tons in 1980 and 29,000 tons in 1990, and for lindane it was 5900 tons in 1980 and 4000 tons in 1990 (Li et al., 1996). HCH has been detected worldwide in all environmental matrices. In particular, HCH has been found in air samples collected in China (Wu et al., 2010), in the Arctic (Becker et al., 2008), and in remote regions such as the Marshall Islands, the Mauritius and southern Chile (Fiedler et al., 2013), just to cite a few examples. In 2009, HCH, both the α- and γ- isomers, as well as lindane, were included in the Stockholm Convention on Persistent Organic Pollutants list, which has led to their phase out. In this paper, we will analyze the atmospheric concentrations of the DDTs and HCHs at five sites around the Great Lakes with an emphasis on the relative isomeric compositions of both DDT and HCH. The purpose of the paper is to report both on the temporal trends of these compounds' concentrations and on the relative amounts of the different DDT congeners and of the different HCH conformers. Our goal is to use these data to identify sources of these compounds and to determine the temporal strengths of these sources. This information will provide insights into the effectiveness of regulatory efforts. 2. Materials and methods 2.1. Sampling and analytical methodology All of the samples were collected as part of the Integrated Atmospheric Deposition Network (IADN), which has been collecting air samples (both the vapor and particle phases) every 12 days at several
Fig. 1. Structures showing the commercial synthesis of DDT and dicofol and the environmental degradation (weathering) of DDT to DDE and DDD. o,p′-DDT is the major impurity (at ~10%) in dicofol.
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Fig. 2. Locations of the atmospheric sampling sites around the North American Great Lakes. The red shaded areas represent population centers.
sites on the shores of the Great Lakes since 1991. The samples were collected at the following United States IADN sites (Fig. 2): Eagle Harbor, Michigan, a remote site on Lake Superior; Sleeping Bear Dunes, Michigan, a very rural site on Lake Michigan; Chicago, Illinois, a highly urbanized site on Lake Michigan; Cleveland, Ohio, an urban site on Lake Erie; and Sturgeon Point, New York, a rural site on Lake Erie (for details, see Environment Canada, 2010). Measurements of these compounds generally started in 1991–1992 at these sites, but the precise timelines over which the measurements were made are given in Fig. S1. Details of the sample collection, extraction, and analysis procedures can be found elsewhere (Salamova and Bayes, 2013; Carlson et al., 2004), and only a brief description will be presented here. A modified Anderson high-volume air sampler (General Metal Works, model GS2310) was used to collect air samples for 24 h at a flow rate giving a total sample volume of about 820 m3. The air stream was first passed through a Whatman quartz fiber filter (QM-A, 20.3 × 25.4 cm) to collect the particles and then through Amberlite XAD-2 resin (Supelco, 20–60 mesh), held in a stainless steel cartridge, to collect the vapor phase components. After sampling, the filters and the XAD were separately Soxhlet extracted for 24 h with 50% (v/v) acetone in hexane. Prior to extraction, the samples were spiked with recovery standards (dibutylchlorendate, δ-HCH, and ε-HCH). The extract was reduced in volume by rotary evaporation, the solvent was exchanged to hexane, and this solution was fractionated on a column containing 3.5% (w/w) water deactivated silica gel. The column was eluted with 25 mL of hexane (fraction 1) and 25 mL of 50% (v/v) hexane in dichloromethane (fraction 2). After N2 blow down, the samples were spiked with internal standards (PCB-65 and PCB-165) that were then used for quantification of the target compounds. The DDTs and HCHs were analyzed by gas chromatography on Agilent Technologies (Palo Alto, CA) 5890 and 6890 instruments equipped with 63Ni electron capture detectors and DB-5 and DB-1701 (J & W Scientific) 60-m columns (250-μm i.d. and 0.1-μm film thickness). Quantitation used the internal standard method. Recovery standards were used to estimate recoveries of each compound in each sample. Results were not recovery corrected because recoveries were in the range 80–120%.
2.2. Statistical calculations All statistical calculations were done with Minitab 16. We calculated ratios of concentrations (see below), which have a higher precision than the concentrations alone. In general, all of these ratios had relative standard deviations of b 10%, which is at least twice as precise as the individual concentration measurements. 2.3. Quality control and quality assurance Quality control and quality assurance procedures were followed to ensure data accuracy. The detailed procedures are described in the IADN Quality Assurance Program Plan and in the IADN Quality Control Project Plan (Basu and Arnold, 2011). 3. Results and discussion The concentrations of these compounds were measured in both the vapor and particle phases for samples collected over different time periods depending on the site and the analyte (see the various sampling timelines in Fig. S1). Using the vapor phase and particle phase data that overlapped in time (for example, p,p′-DDT and α-HCH at Sturgeon Point and Sleeping Bear Dunes for the years 1997–2002, inclusive), we determined that, in general, the fraction of the DDTs in the particle phase was b 15% and that the fraction of α- and γ-HCHs in the particle phase was b3%. These were small amounts, and in fact, this reasoning earlier in the project lead us to abandon the particle phase measurements of pesticide concentrations in May 2003 at all sites except Chicago and Cleveland. For consistency, this paper deals only with the vapor phase concentrations of these compounds. As an additional simplification, the 30 concentrations measured during one year were averaged over that year for each of the compounds in this study. In this way, seasonal effects, which we have shown can be substantial (Venier and Hites, 2010a), were eliminated. These concentration averages are tabulated in the Supporting data and are shown as semilogarithmic plots as a function of time in Figs. S2–S6.
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Many of these concentrations are a function of both the population density at the sampling site and the year of sampling. For example, the atmospheric concentrations of p,p′-DDT at Chicago are much greater than those at Eagle Harbor, and in both cases, the concentrations decrease significantly with time (see Figs. S2 and S6). To model these relationships, we used the following equation to fit the data 2
ln ðC Þ ¼ a0 þ a1 t þ a2 log ðpopÞ
ð1Þ
where t is the sampling year, pop is the number of people living and working within a 25-km radius of the sampling site, a0 is the intercept that rectifies the units, a1 is a rate constant in years−1, and a2 quantitates the effect of population. The time it takes for an atmospheric concentration to decrease by half is called its halving time and is given (in years) by t 1=2 ¼
− ln ð2Þ a1
ð2Þ
The justification for using this model, specifically the population term, has been presented previously by Venier and Hites (2010a,b). The importance of each of the regression parameters in explaining the variability of the concentrations was estimated by calculating the relative contribution of each term to the total variability. This value is reported as the percentage of the sum of squares (SOS) attributed to each factor relative to the total sum of squares. The sum of squares for each term adds up to the model (or regression) sum of squares, allowing us to express the importance of each term as a percent of the overall r2 value. Table 1 gives a summary of the regression results for all of the compounds discussed here. Table S6 gives the complete regression output. 3.1. DDT and related compounds Table 1 shows that the regressions using Eq. (1) are highly significant for all of the DDT-related compounds with F N 47. Four compounds, p,p′-DDT, p,p′-DDE, o,p′-DDT, and ∑DDT, give similar overall regression results. The r2 values average about 86%, the population terms average about 0.071 and represent about 81% of the total sum-ofsquares, and with the exception of o,p′-DDT, the time terms average about −0.082 year−1 and represent 2–9% of the total sum-of-squares. This time term converts to a halving time of 8–9 years. This concordance suggests that these four compounds are behaving similarly to
one another in the environment and that our previous use of ∑ DDT to characterize this behavior was appropriate. The other two compounds of this group, p,p′-DDD and o,p′-DDD, gave regression results that were similar to each other but different than the other four compounds in this group. The r2 values averaged about 65%, the population terms were also lower, at about 0.034 (representing ~ 16% of the total sum-of-squares), and the time terms were much higher, at about −0.21 year−1 (representing about 49% of the total sum-of-squares). This average time term converts to a halving time of 3–4 years. Why do these two DDD-related compounds behave so much differently in the atmosphere compared to the DDT- and DDE-related compounds? The difference in halving times between DDE and the DDD isomers might be related to different DDT degradation pathways: Under aerobic conditions, DDT is degraded to DDE, the most persistent of all DDT related compounds, but under anaerobic conditions it degrades to DDD. The first pathway is the most prevalent and DDE is often resistant to biodegradation, providing a possible explanation for the longer halving times of DDE versus DDD (Ricking and Schwarzbauer, 2012; Spencer et al., 1996; Strompl and Thiele, 1997; Zitko, 2003). In the regressions for p,p′-DDT, p,p′-DDE, o,p′-DDT, and ∑ DDT, almost all of the variability is accounted for in the population term. Fig. 3 shows this part of the regression as the logarithm of the median concentrations as a function of the population term. These are all significant positive relationships, indicating that the median concentrations of these four DDT-related compounds are much higher in the cities of Chicago and Cleveland than at the rural and remote sites. This is surprising given that DDT had been used as an agricultural insecticide, and there has been little agriculture in the cities. It is likely that this urbaneffect was caused by the use of DDT in cities in the 1950s and 1960s for insect control in shade trees that lined many streets at that time. Apparently, DDT is being released from urban soil [or perhaps from treated buildings as suggested by Braeuner et al. (2011) and Dirtu et al. (2012)] into the Great Lakes' atmosphere. This relationship with population ultimately indicates that if the population would decline, the concentration would also decline; at the moment, though, it is not possible to speculate on how long it might take to see this effect since we do not understand environmental hysteresis. One way of determining the sources and origins of DDT in environmental samples is to calculate isomer ratios among DDT and one or more of its environmental degradation products: DDE and DDD. For example, one could calculate the ratio of p,p′-DDE to p,p′-DDT to get an estimate on how long the DDT has been in the environment. This
Table 1 Regression parameters for annual average concentrations of several DDT and HCH-related compounds and their ratios.
Constant term Standard error Probability Time term Standard error Probability %SOS Halving time, t1/2 Standard error, t1/2 Population term Standard error Probability %SOS r2 Prob, r2 N F
p,p′-DDT
p,p′-DDE
o,p′-DDT
∑DDT
p,p′-DDD
o,p′-DDD
Rp,p′
Ro,p′
α-HCH
β-HCH
γ-HCH
Rγ
175 13 b0.1% −0.0874 0.0066 b0.1% 8.9% 7.9 0.6 0.0652 0.0028 b0.1% 79.6% 88.5% b0.1% 81 300.
149 21 b0.1% −0.0745 0.0107 b0.1% 2.2% 9.3 1.3 0.0770 0.0046 b0.1% 76.8% 79.0% b0.1% 81 146
NS NS NS NS NS NS NS NS NS 0.0733 0.0037 b0.1% 86.6% 88.7% b0.1% 54 199
168 15 b0.1% −0.0832 0.0073 b0.1% 6.7% 8.3 0.7 0.0670 0.0031 b0.1% 79.9% 86.6% b0.1% 81 251
320 29 b0.1% −0.160 0.015 b0.1% 48.8% 4.3 0.4 0.0331 0.0063 b0.1% 13.6% 62.4% b0.1% 80 63.8
500. 54 b0.1% −0.250 0.027 b0.1% 48.7% 2.8 0.3 0.0340 0.0066 b0.1% 18.3% 67.0% b0.1% 50 47.8
NS NS NS NS NS NS NS – – 0.00247 0.00075 0.2% 11.8% 14.9% 0.2% 81 6.83
−37.7 5.0 b0.1% 0.0189 0.0025 b0.1% 46.9% – – 0.00253 0.00062 b0.1% 13.0% 59.9% b0.1% 54 38.1
397 12 b0.1% −0.196 0.006 b0.1% 93.3% 3.5 0.1 NS NS NS NS 93.6% b0.1% 81 569
443 42 b0.1% −0.221 0.021 b0.1% 49.2% 3.1 0.3 0.0393 0.0051 b0.1% 28.3% 77.5% b0.1% 51 82.5
407 17 b0.1% −0.202 0.008 b0.1% 78.9% 3.4 0.1 0.0285 0.0035 b0.1% 9.6% 88.5% b0.1% 81 301
NS NS NS NS NS NS NS – – 0.00645 0.00040 b0.1% 70.8% 78.8% b0.1% 81 145
The ratios Rp,p′, Ro,p′, and Rγ are defined by Eqs. (3), (5), and (7). The regression parameters for the three terms were determined from Eq. (1), (4), (6), or (8) as appropriate. Halving times (t1/2, in years) were calculated using Eq. (2). See the Supporting data for data plots and for detailed regression results. “NS” means not significant at P b 1%.
M. Venier, R.A. Hites / Environment International 69 (2014) 159–165
Median concentration (pg/m3)
100
ΣDDT (0.809) 10
p,p'-DDE (0.751) 1
p,p'-DDT (0.860) o,p'-DDT (0.974) 0.1 0
10
20
30
40
50
log2 (pop) Fig. 3. Median vapor-phase concentrations of p,p′-DDT, p,p′-DDE, and o,p′-DDT, and ∑DDT as a function of the squared logarithm of the number of people living and working within a 25-km radius of the sampling site. ∑DDT is the sum of the p,p′-DDT, p,p′-DDE, and p,p′-DDD concentrations. The numbers in parentheses are the r2 values for each regression; all are significant at P b 6%.
environmental degradation process is frequently called “weathering.” In this case, we will use the following unitless ratio:
Rp;p′
p; p′‐DDE þ p; p′‐DDD ¼ p; p′‐DDE þ p; p′‐DDD þ p; p′‐DDT
ð3Þ
where [p,p′-DDT], [p,p′-DDD], and [p,p′-DDE] are the concentrations (in pg/m3) of each isomer in the atmospheric vapor phase. This ratio has been normalized, compared to the one commonly used in the literature (Li et al., 2007; Liu et al., 2009; Munoz-Arnanz and Jimenez, 2011; Park et al., 2011; Qiu et al., 2004; Wang et al., 2007; Yang et al., 2008), to set its range between 0 and 1, rather than between 0 and infinity. These ratios were calculated for each site and for each sampling year, and are all given in the Supporting data. To investigate the effect of time and of population density around the sampling site, the following regression was used 2
Rp;p′ ¼ b0 þ b1 t þ b2 log ðpopÞ
ð4Þ
0.8
163
where the bi parameters correspond to the ai parameters in Eq. (1). The regression parameters are given in Table 1. This analysis showed a small, but significant, overall correlation (r2 = 14.9%), and the only significant term was the one associated with population. This part of the regression is plotted in Fig. 4. Given this lack of a strong relationship with time or population, we have simply taken the average of all Rp,p′ values across all five sites, and this average is 0.648 ± 0.011. This is similar to values in the literature for air samples collected near Lake Taihu in China (0.48–0.76) (Qiu et al., 2004) and for air samples collected in remote southwestern China (0.65 ± 0.39) (Xu et al., 2011). Ratios suggestive of dicofol-type DDT sources were reported in the Arctic (Becker et al., 2012; Bidleman et al., 2013). There are no precise guidelines on how to interpret Rp,p′ values, but higher values suggest the presence of significant weathering of DDT. An apparent global convergence of Rp,p′ values at 0.6–0.7 may indicate that a rough equilibrium between the rate of degradation of p,p′-DDT to p,p′-DDE and p,p′-DDD and their subsequent degradation (perhaps to monocyclic compounds) has been reached. It is also possible that there is an environmental source of DDT other than technical DDT, and that source could be dicofol (Qiu et al., 2005). One way to distinguish DDT inputs from dicofol inputs is to assess the concentrations of o,p′-DDT, which is the most abundant impurity in technical grade dicofol, relative to the concentrations of p,p′-DDT itself (Qiu et al., 2005). Thus, we define another ratio to compare the relative abundances of the o,p′- to p,p′-DDT isomers: o; p′‐DDT Ro;p′ ¼ o; p′‐DDT þ p; p′‐DDT
ð5Þ
where o,p′-DDT and p,p′-DDT are the concentration of each isomer in the vapor phase (pg/m3). As before, we investigated the effect of time and the site's population density using the following regression 2
Ro;p′ ¼ c0 þ c1 t þ c2 log ðpopÞ
ð6Þ
where the ci parameters correspond to the ai parameters in Eq. (1). The resulting regression parameters are given in Table 1. In this case, all three of the fitted parameters are significant, and the overall r2 value is 60%, which indicates a strong relationship. Of the total sum-ofsquares, most (47%) is from the time-term, and this relationship is shown specifically in Fig. 5. This observation is different than that observed for Rp,p′, which showed no significant temporal variations. This increase in the relative abundance of the o,p′-DDT with passing year is unusual, and warrants a bit more examination. Can we distinguish the situation in which the absolute concentration of dicofol has
0.6 0.5
Rp,p' (0.069)
0.4
Average Ro,p'
Average Rp,p' or Rγ
0.6
0.2
Rγ (0.981)
0.4 0.3 0.2
Ro,p' (0.699)
0.1
0.0 0
10
20
30
40
50
log2 (pop) Fig. 4. Relationship between the overall average Rp,p′ and Rγ values (see Eqs. (3) and (7) for definitions of these ratios) at each site and the population living and working within 25 km of that site. The standard errors of each average are smaller than the symbols. The numbers in parentheses are the r2 values for each regression. The Rp,p′ regression is not statistically significant, but that for Rγ is significant at P b 0.1%.
0.0 1998
2000
2002
2004
2006
2008
2010
2012
Fig. 5. Annual averages of Ro,p′ (see Eq. (2)) in air sampled near the Great Lakes as a function of sampling year. The averages were over the five sites for each year. The error bars are standard errors. The number in parentheses is the r2 value of the regression, which is significant at P b 0.2%.
164
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increased as opposed to the situation in which the relative concentration of dicofol has increased? The temporal data for o,p′-DDT shown in Figs. S2–S6 show that there is no relationship between the concentrations of this compound and time, but there is a strong relationship between the concentrations of p,p′-DDT and time. These observations suggest that the changes in Ro,p′ are driven by the concentration of p,p′-DDT in the denominator of Eq. (5) and that the amount of dicofol in the atmosphere (as indirectly measured by o,p′-DDT) has remained relatively constant over time while the amount of p,p′-DDT has decreased more rapidly. Because the vapor pressures of o,p′-DDT and p,p′-DDT are different, it is not possible to interpret the environmental value of Ro,p′ in terms of the relative inputs of DDT and dicofol, but it is clear from the rate at which Ro,p′ has increased (as represented by the positive slope in Fig. 5) that the relative concentration of dicofol in Great Lakes air has about doubled over the last decade. This phenomenon has also been observed for Taihu Lake in China (Qiu et al., 2004). Very little data on air measurements of dicofol exists. It was detected in the Arctic (AMAP, 2009) and in samples collected on a cruise from East Asia to the High Arctic (Zhong et al., 2011), hinting that long range transport might be playing a role in the observed levels. Because dicofol was retired from the United States market in 2011, we expect to see a slow decrease of Ro,p′ back to values closer to 0.2 over the next several decades. In fact, the two most recent data points in Fig. 5, corresponding to years 2009 and 2010, are below the regression line, suggesting that this trend may have already started; of course, more data are needed to confirm this suggestion. 3.2. Hexachlorocyclohexane (HCH) The use of this insecticide has long been eliminated. Nevertheless, HCH isomers are still detected in air samples from around the Great Lakes, but perhaps as a result of the restrictions put on their use, these concentrations are now decreasing rapidly. The resulting regression parameters for the three HCH conformers in our dataset are given in Table 1. In this case, the time terms are significant for all three conformers. The overall r2 values are N77%, which indicates a strong relationship, and most of the total sums-of-squares are from the time term. From this term, halving times for α-, β-, and γ-HCH are calculated to be 3.1–3.5 years. This agrees with results we have reported before for α- and γ-HCH (Venier and Hites, 2010a), but results for β-HCH have not been reported before. There are no significant differences in the halving times among the three compounds. For β- and γ-HCH, the population terms are significant, averaging at 0.034, and represent 10–28% of the total sum-of-squares. Given that both of these terms are positive, it is likely that there are some urban sources of β- and γ-HCH, but not of α-HCH, to the environment. As with DDT, the ratio of the HCH isomers can be used to infer useful information regarding the sources and environmental behavior of these compounds. Since α- and γ-HCH were the major components of the HCH technical mixture, we calculated the unitless fraction of the γ isomer Rγ, which is defined as Rγ ¼
½γ‐HCH ½α‐HCH þ ½γ‐HCH
ð7Þ
where [α-HCH] and [γ-HCH] are the concentrations (pg/m3) measured in the atmospheric vapor phase at each of the IADN sites. Like the previous ratios, we modified the equation from the one that is usually reported in the literature (Qiu et al., 2004; Willett et al., 1998; Xu et al., 2011) to limit its range to between 0 and 1, rather than between 0 and infinity. As before, we investigated the effect of sampling year and the site's population density using the following regression 2
Rγ ¼ d0 þ d1 t þ d2 log ðpopÞ
ð8Þ
where the di parameters correspond to the ai parameters in Eq. (1). The resulting regression parameters are given in the rightmost column of Table 1; the only significant parameter was the one for population. This effect is shown in Fig. 4. Clearly, there is a strong positive relationship between the Rγ value at a given site and the number of people living or working within a 25-km radius of that site. This relationship has not been observed before. We have reported previously that the atmospheric concentrations of neither α-HCH nor γ-HCH were significantly dependent on the local human population (Venier and Hites, 2010b), but the results reported here indicate that Rγ increases as population near the sampling site increases. This seeming conundrum may be the result of comparing absolute concentrations to relative concentrations. These measurements of Rγ are more sensitive to human population effects than the absolute measurements of the concentrations themselves, particularly once the Rγ values are averaged over all years. This finding suggests that technical HCH (high in α-HCH) was used almost exclusively as an agricultural insecticide, but that lindane (high in γ-HCH) may have had some urban uses. The value of Rγ has been reported to be in the range of 0.12 to 0.25 for technical HCH (Willett et al., 1998; Xu et al., 2011), and this is about what we observe at the remote site of Eagle Harbor. 4. Conclusions The concentrations of DDT and related compounds are decreasing in the atmosphere around the North American Great Lakes, but there is some evidence that the DDT found as an impurity in dicofol is now contributing to some of this atmospheric load. There is also some evidence that DDT levels in cities are higher than previously suspected. The atmospheric concentrations of α-, β-, and γ-HCH (lindane) are decreasing rapidly, but there is also evidence that the concentrations of lindane are higher in the cities than previously suspected. The overall implications are that compounds such as DDT and lindane, which were previously thought to have exclusively agricultural uses, may also have been used in cities for insect control and that cities may be significant sources of these compounds to the atmosphere. Acknowledgments We thank the entire Team IADN at Indiana University for sample and data acquisition and the U.S. Environmental Protection Agency's Great Lakes National Program Office for funding (Grant No. GL-00E76601-3, Todd Nettesheim, project officer). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.envint.2014.03.028. References Agency for Toxic Substances and Disease Registry. Toxicological profile for DDT, DDE, and DDE; 2002 [September]. Arctic Monitoring and Assessment Programme (AMAP). Arctic pollution 2009; 2009 [Oslo]. Basu I, Arnold K. Integrated Atmospheric Deposition Network Quality Assurance Project Plan. Bloomington, IN: Indiana University; 2011 (http://www.epa.gov/greatlakes/ monitoring/air2/iadn/qapp/iadn-qapp-201111.pdf). Becker S, Halsall CJ, Tych W, Kallenborn R, Su Y, Hung H. Long-term trends in atmospheric concentrations of alpha- and gamma-HCH in the Arctic provide insight into the effects of legislation and climatic fluctuations on contaminant levels. Atmos Environ 2008;42(35):8225–33. Becker S, Halsall CJ, Tych W, Kallenborn R, Schlabach M, Mano S. Changing sources and environmental factors reduce the rates of decline of organochlorine pesticides in the Arctic atmosphere. Atmos Chem Phys 2012;12:4033–44. Bidleman TF, Kurt-Karakus PB, Wong F, Alegria HA, Jantunen LM, Hung H. Is there still “new” DDT in North America? An Investigation using proportions of DDT compounds. In: McConnell LL, Dachs J, Hapeman CJ, editors. Occurrence, Fate and Impact of Atmospheric Pollutants on Environmental and Human Health, vol. 1149. Washington: Amer Chemical Soc; 2013. p. 153–81.
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