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Keywords:

  • Aquatic risk;
  • Estrogens;
  • Environmental modeling;
  • PhATE™ mixtures

Abstract

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSION
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

In an effort to assess the combined risk estrone (E1), 17β-estradiol (E2), 17α-ethinyl estradiol (EE2), and estriol (E3) pose to aquatic wildlife across United States watersheds, two sets of predicted-no-effect concentrations (PNECs) for significant reproductive effects in fish were compared to predicted environmental concentrations (PECs). One set of PNECs was developed for evaluation of effects following long-term exposures. A second set was derived for short-term exposures. Both sets of PNECs are expressed as a 17β-estradiol equivalent (E2-eq), with 2 and 5 ng/L being considered the most likely levels above which fish reproduction may be harmed following long-term and short-term exposures, respectively. A geographic information system-based water quality model, Pharmaceutical Assessment and Transport Evaluation (PhATE™), was used to compare these PNECs to mean and low flow concentrations of the steroid estrogens across 12 U.S. watersheds. These watersheds represent approximately 19% of the surface area of the 48 North American states, contain 40 million people, and include over 44,000 kilometers of rivers. This analysis determined that only 0.8% of the segments (less than 1.1% of kilometers) of these watersheds would have a mean flow E2-eq concentration exceeding the long-term PNEC of 2.0 ng/L; only 0.5% of the segments (less than 0.8% of kilometers) would have a critical low flow E2-eq exceeding the short-term PNEC of 5 ng/L. Those few river segments where the PNECs were exceeded were effluent dominated, being either headwater streams with a publicly owned treatment works (POTW), or flowing through a highly urbanized environment with one or several POTWs. These results suggest that aquatic species in most U.S. surface waters are not at risk from steroid estrogens that may be present as a result of human releases. Environ. Toxicol. Chem. 2012;31:1407–1415. © 2012 SETAC


INTRODUCTION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSION
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

The endogenous and synthetic steroid estrogens (estrone [E1], 17β-estradiol [E2], estriol [E3], 17α-ethinyl estradiol [EE2]) appear to be the most estrogenic component of sewage effluent 1–3). They have been detected in publicly owned treatment works (POTW) effluents and surface waters on every continent including the following: North America, such as the United States 4 and Canada 5; South America, such as Brazil 6; East Asia, including China 7 and Japan 8; Australia 9; Africa, such as South Africa 10; Europe, including the United Kingdom 11, Switzerland 12, The Netherlands 13, France 14; Italy 15; and the Middle East 16. Their presence in surface waters at low ng/L concentrations is consistent with modeling results that also predict their presence as a result of human releases 17–19. Laboratory studies have shown that steroid estrogens can cause a variety of effects in aquatic biota (summarized in Caldwell et al. 20, 21). Controlled field investigations have also reported adverse effects, although these occurred at concentrations substantially higher than are expected to occur in natural settings 22. Disruption to the reproductive capabilities of male fish may ultimately put at risk the success of resident fish populations 22, 23.

In rivers, effects assumed to be related to exposure to estrogens, or compounds with estrogenic activity, have been reported downstream of water courses receiving sewage effluent 9, 13, 24–29. In the United Kingdom, although it has been postulated that animal husbandry may represent a significant source of estrogens to the environment 30, 31, the clearest link to the degree of endocrine disruption in wild fish has been with the proportion of sewage effluent in the river 29. What appear to be estrogen-related effects (e.g., changes in sex ratio, vitellogenin [VTG] induction) have also been reported downstream of POTWs in the United States 32–35. However, the importance of human-derived steroid estrogens relative to other chemicals or factors in U.S. surface waters as the cause of these observed effects remains unclear 32, 34, 36.

As a highly developed country with a large and growing population, it is clearly important to assess the level of endocrine disruption risk to fish across the United States from human-derived steroid estrogens. Assessing risk to natural fish populations across regions, or nations, can be attempted using large-scale monitoring programs or geographic modeling exercises. As discussed in Johnson et al. 37, both have their supporters and detractors, strengths and weaknesses, but where measured environmental concentrations (MECs) and model predicted environmental concentrations (PECs) show acceptable agreement, we can have a great deal more confidence in the conclusions. Recently, geographic information system (GIS)-based water quality models have been successfully used to predict endocrine disruption risk across the United Kingdom 19 and Switzerland 38. The present study has chosen to follow this approach using the Pharmaceutical Assessment and Transport Evaluation (PhATE™) model 17 for 12 U.S. watersheds combined with a critical review of the aquatic toxicity of the steroid estrogens 20, 21 to characterize the potential endocrine disruption risk to U.S. surface waters.

A critical issue, as raised by Johnson et al. 37, is how accurate do the model predictions of estrogen concentrations need to be? If the predicted concentrations are considerably below levels associated with potential risk, then small errors or minor oversimplifications in the modeling exercise will be irrelevant. Anderson et al. 17 previously used MECs reported for several sewage-derived compounds to corroborate the PECs estimated by PhATE for 11 of these U.S. watersheds. (The Raritan River has since been added to PhATE, increasing the total number of watersheds included in the model to 12.) Similarly, Hannah et al. 18 extended those findings for EE2, demonstrating that PECs predicted by PhATE are consistent with MECs generated using sensitive analytical methods. Thus, there is evidence that the PhATE model is capable of giving a reasonable estimate of the potential risk to fish reproduction from human-derived estrogens in the U.S. freshwater environment.

The present study had four major objectives. The first objective was to select predicted-no-effect concentrations (PNECs) protective of fish reproductive effects during long-term and short-term exposures to the combined major steroid estrogens. The second objective was to use the PhATE model to predict steroid estrogen discharge to the U.S. aquatic environment based on human therapeutic use (for the synthetic estrogen, EE2, and hormone replacement therapy [HRT] and naturally produced estrogens), excretion rates, and fate and behavior information. In the present analysis, conjugated equine estrogens from HRT (i.e., equilin and equilenin) were not directly considered because of limited information on their treatment removal, environmental fate, and potential toxicity. However, an allowance has been made for HRT products that would lead to the additional excretion of E1 and E2 (see Supplemental Data). The third objective was to predict steroid estrogen concentrations as combined 17β-estradiol equivalents (E2-eq) in 12 U.S. watersheds using the PhATE model under mean and critical low flow conditions. The final objective was to provide an overall assessment of the degree of risk to reproduction in fish from human-derived steroid estrogens across the watersheds of the United States by comparing predicted environmental concentrations to long- and short-term PNECs.

METHODS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSION
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

Selection of a PNEC

Selecting a PNEC for the combined major estrogens that would be protective of fish reproductive effects is a different approach from selecting a level below which no effects whatsoever are assumed. Although it is accepted that subtle endocrine disruption effects, such as VTG induction, may occur at concentrations below those where reproductive effects could start, it is fish reproductive ability that is of greatest concern.

Caldwell et al. 20, 21 used species sensitivity distributions of available chronic fish reproductive toxicity data to develop PNECs of 0.1 and 2.0 ng/L for EE2 and E2, respectively, protective of reproductive effects in fish following long-term exposures (i.e., exposures encompassing several life stages or multiple generations). These PNECs were used in the present study. Unfortunately, insufficient data were available to employ the same methodology to derive PNECs for E1 and E3 21. Instead, Caldwell et al. 21 employed in vivo VTG induction studies to determine the relative ability to induce VTG by each of the steroid estrogens on the assumption that the estrogens would have the same relative effects on reproductive end points. Thus, Caldwell et al. 21 derived PNECs protective of long-term exposures of 6 and 60 ng/L for E1 and E3, respectively.

Caldwell et al. 21 recognized that E2 and EE2 are unique in that the potential aquatic effects of each has been investigated several times by studies that expose fish to constant levels of these steroid estrogens for multiple life stages and even multiple generations. Such constant long-term exposure seldom, if ever, occurs in natural waters. Concentrations are expected to vary, being lowest during high river flows and highest during low river flows. It is common for aquatic risk assessments to focus on the effects of the highest concentrations that occur during periods of low river flow by comparing low flow PECs to PNECs. By definition, such critical low flows only last for several days, not the many months used in the multi-generation studies to derive the long-term PNECs. Flows closer to the critical low flow than the mean annual flow may be present for several weeks or months during, for example, the late summer and fall in many U.S. rivers; however, those low flows typically do not persist for the entire life cycle of a fish species.

Caldwell et al. 21 found that the no-observed-effect concentrations (NOECs) for E1 and E2 reported by studies investigating exposures of less than 60 d were higher than the NOECs from long-duration, multi-generation studies. The short-term studies developed NOECs for reproductive effects such as changes in sex ratio, intersex, fecundity, and fertilization. In several instances, VTG induction, which is a potential measure of transitory exposure to estrogens, was observed at concentrations that did not elicit the more serious reproductive effects for which NOECs were reported and on which the short-term PNECs are based. On that basis, Caldwell et al. 21 derived PNECs protective of short-term exposures (<60 d) of 20, 5, 200, and 0.5 ng/L for E1, E2, E3, and EE2, respectively.

Two sets of PNECs are used in this risk characterization to evaluate the potential aquatic risks associated with exposure to steroid estrogens. Potential aquatic risks associated with long-duration exposures (represented by mean annual flow PECs) are characterized using PNECs of 6, 2, 60, and 0.1 ng/L for E1, E2, E3, and EE2, respectively. Potential aquatic risks associated with short-term exposures (<60 d) are assessed using PNECs of 20, 5, 200, and 0.5 ng/L for E1, E2, E3, and EE2, respectively. For the risk assessment, these estrogens were assumed to act in an additive way so that after accounting for their relative potencies, their concentrations in surface water could be summed to derive an E2-eq concentration 40, 41.

The PECs of estrogens in U.S. surface waters

Mean and critical low flow (i.e., 7Q10) PECs of estrogens in surface water resulting from humans were estimated using the PhATE model, Ver. 2.1.1 17. This GIS-based model is able to generate PECs for 12 watersheds in the United States. The watersheds included in PhATE were selected to represent a typical range of watersheds in the United States 17. The watersheds (1,488,661 km2) make up about 19% of the land area and were selected to represent the range of geography and climates present in the contiguous 48 states. They include 1,132 POTWs on about 41,508 kilometers of rivers divided into 2,713 segments. Fifty-seven of the POTWs included in PhATE serve over 50,000 people. In total, 692 have secondary treatment, 391 have advanced treatment, 7 have either primary or advanced primary treatment, and 42 do not discharge to a surface water. About 40 million people live within the 12 watersheds. The hydrological data underpinning the model are derived from U.S. Geological Survey river flow measurements, collected over varying durations, depending on watersheds through 2000.

The PhATE model requires several compound-specific inputs including the per capita use for prescribed estrogens or excretion rate for naturally occurring estrogens, metabolism for prescribed estrogens, POTW removal rate, and in-stream removal rate. Key inputs are summarized in Table 1 and are discussed in more detail in the Supplemental Data available online. The total amount of estrogens released into surface water through POTWs as a result of humans was estimated by summing the excreted mass of estrogens sold for therapeutic use in the United States and the excreted mass of naturally produced estrogens (Table 1 and Supplemental Data). Note that nonpoint sources such as releases from animal husbandry operations are not included in the present study. Excreted mass of estrogens from therapeutic use was estimated by adjusting annual sales volume by reported metabolism. Excreted mass of naturally produced estrogens was estimated by identifying age- and gender-specific excretion rates and weighting those rates according to the fraction of the total population that each age and gender category comprises in a method similar to Johnson and Williams 42 (see Supplemental Data). Assumed POTW removal rates are the median of removal rates reported in the literature (Table 1 and Supplemental Data). In-stream decay rates, for which relatively few studies are available, are based on on a review of the literature (Table 1 and Supplemental Data). Note that in the portion of a river reach immediately downstream of a POTW outfall, little degradation occurs because residence times for discharged estrogens are very short.

Table 1. Summary of key inputs used in the Pharmaceutical Assessment and Transport Evaluation Model to estimate predicted environmental concentrationsa
CompoundMetabolism (%)bTotal mass excretedPublicly owned treatment works removalIn-stream decay (1/d)
Per capita (µg/d)Total U.S. (kg/y)cSecondary (%)Advanced secondary (%)
  • a

    Additional detail regarding the derivation of the inputs summarized in this table are provided in the Supplemental Data.

  • b

    Metabolism only used to estimate excreted mass of prescribed hormone.

  • c

    Total mass excreted assumes a U.S. population of 277,048,382 (U.S. Census, 2001).

Estrone80191,80267880.3
17β-Estradiol607.778085960.3
Estriol0818,20697975.7
17α-Ethinyl Estradiol500.4141.284840.07

Estimating E2-eq concentrations

Within each river segment modeled by PhATE, PECs of each individual estrogen were converted to an E2-eq concentration and then the E2-eq concentrations of each individual estrogen were summed to estimate a total E2-eq concentration for each river segment. The 17β-estradiol equivalents concentrations of each individual estrogen were derived by multiplying the concentration of each estrogen by its relative estrogenic potency to E2. Relative potency of the different estrogens was based on the relative difference of PNECs for each of the estrogens 21. For example, for long-term exposures, EE2 was assumed to have a relative potency of 20 (equal to the E2 PNEC [2 ng/L] divided by the EE2 PNEC [0.1 ng/L]). Similarly, E1 and E3 were assumed to have relative potencies of 0.33 and 0.033, respectively 21, for long-term exposures. For short-term exposures, the relative potencies were assumed to be 0.25, 0.025, and 10 for E1, E3, and EE2, respectively. Thus, a river reach that has mean flow concentrations of 0.1, 0.02, 0.004, and 0.004 ng/L of E1, E2, E3, and EE2, respectively (equal to the ∼90th percentile mean flow PECs; Supplemental Data) was assumed to have a long-term E2-eq concentration of 0.13 ng/L ([0.1 ng/L E1 × 0.33] + 0.02 ng/L E2 + [0.004 ng/L E3 × 0.033] + [0.004 ng/L EE2 × 20]). The short-term E2-eq concentration for those same steroid estrogen concentrations (equal to between the 60th to 90th percentile of critical low flow PECs depending on steroid estrogen; Supplemental Data) would be 0.09 ng/L ([0.1 ng/L E1 × 0.25] + 0.02 ng/L E2 + [0.004 ng/L E3 × 0.025] + [0.004 ng/L EE2 × 10]). The E2-eq concentrations were estimated for all 2,713 reaches in the 12 watersheds modeled by PhATE. Annual average flow PECs were used to develop long-term E2-eq concentrations and critical low flow concentrations were used to estimate short-term E2-eq concentrations. By converting all four steroid estrogens into E2-eq and combining their E2-eq concentrations, this evaluation inherently assumes additivity of the effects of the four estrogens, as has been demonstrated experimentally 40, 41.

Given the recent report of endocrine disruption effects in bass 36, a further analysis was carried out that compared the E2-eq PECs from PhATE and the intersex occurrence and VTG induction data reported by Hinck et al. 36 to investigate whether a relationship existed between PECs and the intersex and VTG observations. Twenty-one of the river reaches for which Hinck et al. 36 reported intersex occurrence in bass overlapped one or more segments in PhATE (Supplemental Data). Sixteen segments for which VTG induction was reported by Hinck et al. 36 overlapped one or more PhATE segments (Supplemental Data). Overlap occurred in the following four watersheds: Mississippi Headwaters, Columbia River, Lower Colorado Basin, Apalachicola River (Supplemental Data). Three reaches for which intersex information were available, and four for which VTG data were available, were represented by more than one PhATE segment. In those cases, the arithmetic average of the E2-eq PECs generated by PhATE for each of the segments was used in the comparison (Supplemental Data).

To examine the effect of uncertainty in the PEC to PNEC comparison, the margin of safety (MOS) was derived and plotted for each PhATE segment in each watershed, as well for all watersheds combined. The MOS is equal to the reciprocal of the PEC:PNEC ratio. Thus, a segment with an annual mean flow E2-eq PEC of 0.4 ng/L has a PEC:PNEC ratio of 0.2, assuming the long-term E2-eq PNEC is 2 ng/L, and a MOS of 5 (PNEC/PEC = 5). The MOS of 5 indicates that the PEC could be up to five times higher than estimated (i.e., it could have been underestimated by as much as fivefold), and still not exceed the PNEC. Alternatively, the PNEC could also be five times lower and still not be exceeded by the PEC. Thus, larger margins of safety have associated with them greater confidence that potential aquatic risks are not present.

RESULTS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSION
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

To evaluate the potential for concentrations of estrogens in surface water to cause adverse effects in aquatic biota, E2-eq PECs were calculated for every river segment in the 12 watersheds modeled by PhATE. The first comparison was of mean flow PECs to the long-term PNEC (Fig. 1) and the second comparison was of low flow PECs to the short-term PNEC (Fig. 2). Under mean flow conditions, only the Lower Colorado, Trinity, Mississippi, and Columbia Rivers have a few segments (6.6, 2.6, 0.3, and 0.1%, respectively) with E2-eq PECs greater than the long-term PNEC of 2.0 ng/L (Table 2, Fig. 1). These results indicate that adverse effects on fish reproduction are not expected at mean flow in the vast majority of U.S. surface waters because mean flow PECs are less than the long-term PNEC. Under critical low flow conditions, 6 of 12 watersheds have at least one segment with a PEC exceeding the short-term PNEC, with the Lower Colorado (2.2%), Kansas (1.5%), and Apalachicola (0.5%) Rivers having the highest percentage of segments exceeding the PNEC (Table 2, Fig. 2).

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Figure 1. The cumulative distribution of U.S. mean flow predicted environmental concentrations (PECs) for human-derived steroid estrogens expressed as 17β-estradiol (E2) equivalents are plotted for each individual watershed. The PECs for each individual watershed are shown as a unique color (see legend in the figure). The PECs are also plotted for all watersheds combined (black line). The long-term E2 predicted-no-effect concentration (PNEC) of 2 ng/L is shown as a horizontal solid red line.

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Figure 2. The cumulative distribution of U.S. low flow predicted environmental concentrations (PECs) for human-derived steroid estrogens expressed as 17β-estradiol (E2) equivalents are plotted for each individual watershed. The PECs for each individual watershed are shown as a unique color (see legend in the figure). The PECs are also plotted for all watersheds combined (black line). The short-term E2 predicted-no-effect concentration (PNEC) of 5 ng/L is shown as a horizontal solid red line.

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Table 2. Summary of the percentage of segments and river kilometers with a mean flow E2-eq PECa greater than the long-term PNEC of 2 ng/L and a low flow E2-eq PECb greater than the short-term PNEC of 5 ng/L
WatershedLow flow PEC > 5 ng/LMean flow PEC > 2 ng/L
NameKilometersSegments% Kilometers% Segments% Kilometers% Segments
  • a

    E2-eq estimated assuming PNECs of 0.1, 2, 6, and 60 ng/L for EE2, E2, E1, and E3, respectively.

  • b

    E2-eq estimated assuming short-term PNECs of 0.5, 5, 20, and 200 ng/L for EE2, E2, E1, and E3, respectively. E2-eq = 17β-estradiol equivalents; PEC = predicted environmental concentration; PNEC = predicted-no-effect concentration; EE2 = 17α-ethinyl estradiol; E1 = estrone; E3 = estriol.

Merrimack River644470.00.00.00.0
Little Miami River, Ohio1,304740.00.00.00.0
White River, Indiana3,0581190.00.00.00.0
Mississippi Headwaters11,7215860.20.20.20.3
Kansas River6,0543412.21.50.00.0
Trinity River3,3182240.20.43.22.6
Lower Colorado Basin3,8751794.02.28.86.6
Sacramento River1,164590.00.00.00.0
Columbia River9,3998550.30.30.20.1
Atlanta Headwaters      
 Apalachicola River2,9211891.40.50.00.0
Schuykill River501170.00.00.00.0
Raritan River216230.00.00.00.0
Total44,1742,7130.80.51.10.8

On further examination, segments with an E2-eq PEC exceeding either the long-term or short-term E2 PNEC were almost always headwaters with a POTW, or segments that are flowing through, or have just flowed through, an urbanized portion of a watershed. A typical example of these conditions is the Trinity River Basin (Fig. 3) where segments with locations with PECs exceeding the PNEC can be examined. Under low flow conditions, the E2-eq PECs in only one segment exceeded the short-term PNEC of 5 ng/L. The segment is just south of Dallas/Fort Worth and is a headwater with a POTW. All six segments with a mean flow E2-eq PEC greater than the long-term E2 PNEC of 2 ng/L occur in the vicinity of Dallas/Fort Worth, a major metropolitan area (Fig. 3).

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Figure 3. The 17β-estradiol equivalents (E2-eq) predicted environmental concentrations (PECs) for the Trinity River segments included in the Pharmaceutical Assessment and Transport Evaluation (PhATE) model are shown. Segments for which PhATE estimates PECs are highlighted in blue and locations of sewage treatment plants are shown as open squares. The segment with E2-eq PECs exceeding the short-term E2 predicted-no-effect concentration (PNEC) assuming low flow is highlighted in purple, and the segments exceeding the long-term E2 PNEC assuming mean flow are highlighted in black.

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When PEC exceedances are expressed on percentage of kilometer basis, the general results are similar to when they are expressed on a percentage of segment basis. However, the percentage of kilometers exceeding a PNEC is generally greater than the percentage of segments exceeding the PNEC (Table 2). It should be noted that the percentage of a watershed exceeding a PNEC tends to be an overestimate because of how PhATE assigns PECs to each segment. The PhATE model calculates the mass of a compound present at the end of a river segment and divides that by the river flow at the end of the segment to derive a PEC. The model then assumes the entire segment has the same PEC as estimated for the end of the segment, upstream as well as downstream of a POTW discharge. Such an oversimplification can be seen in the influence of the POTW on most of the urbanized segments in Dallas/Fort Worth (Fig. 3).

Given the recent report of endocrine disruption effects in bass 36, a further analysis was carried out to match 21 of the river reaches (in the Mississippi Headwaters, Columbia River, Lower Colorado Basin, Apalachicola River) in which Hinck et al. 36 measured occurrence of intersex with river reaches modeled in PhATE (Fig. 4; Supplemental Data). Though insufficient data are available to examine differences between watersheds, when the data for all watersheds are combined, no relationship between long-term E2-eq PECs and either the incidence of intersex fish (Fig. 4) or the plasma concentration of VTG (Supplemental Data) was apparent.

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Figure 4. Scatterplot showing the relationship between the frequency of intersex bass (in percent) reported by Hinck et al. 36 and the 17β-estradiol equivalents (E2-eq) predicted environmental concentrations (PECs; in ng/L) estimated by the Pharmaceutical Assessment and Transport Evaluation (PhATE) model for all segments in the United States where paired information was available.

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Finally, when the MOS for every segment included in PhATE was plotted (assuming mean annual flow E2-eq PECs and a long-term PNEC of 2 ng/L; Fig. 5), most have high MOS with fewer than 3% of segments in the United States predicted to have an MOS of 3 or less, and more than 92% have an MOS of greater than 10 (Fig. 5). Under low flow and using the short-term E2 PNEC, approximately 94% of segments have a MOS greater than 3 and approximately 80% have a MOS greater than 10.

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Figure 5. Margins of safety (MOS) for 17β-estradiol (E2) equivalents compared to the long-term E2 predicted-no-effect concentration (PNEC) of 2 ng/L are plotted for each individual watershed assuming mean flow conditions. The MOS for each individual watershed is shown as a unique color (see legend in the figure). The MOS are also plotted for all watersheds combined. Combined United States wide mean flow MOS are shown as a solid black line. Combined United States wide critical low flow MOS (derived using a short-term E2 PNEC of 5 ng/L) are shown as a black dashed line. PEC = predicted environmental concentrations.

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DISCUSSION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSION
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

Risk of exceeding the E2-eq PNEC

Where reproduction effects in biota have been reported, these have generally been the result of long-term, multi-generational studies using a constant estrogen concentration (fish being the most sensitive taxa, and reproduction being the most sensitive endpoint affected by estrogens 20, 21). The finding that E2-eq PECs are greater than the long-term E2 PNEC in only 0.8% of stream segments in the United States under mean flow conditions (Fig. 1) suggests that continuous concentrations of E2-eq greater than 2 ng/L will be rare. Therefore, the severe population level effects reported in laboratory studies and in experimental field situations 22 are not predicted to be widespread in U.S. surface waters.

Flow, and therefore concentration, is not constant and, thus, can be higher or lower than mean flow. Whether adverse reproductive effects occur if fish are exposed to temporary high concentrations of estrogens is not known, though comparison of critical low flow E2-eq PECs to the short-term PNEC of 5 ng/L provides an assessment of this potentiality. Comparison to low flow PECs, which may give rise to short-term high concentration spikes, may be considered appropriate to evaluate the potential for rapid response effects, such VTG induction, frequency of intersex, or changes in gonadosomatic index. The E2 equivalents PECs during critical low flow conditions were also below the short-term E2 PNEC in all but 0.5% of the modeled segments (Table 2, Fig. 2), indicating that more severe reproductive effects associated with exposure to steroid estrogens derived from humans and released from sewage treatment plants are also not expected in most U.S. rivers.

It is likely that flows closer to the critical low flow (7Q10) than the mean annual flow may exist in many U.S. rivers for many weeks to several months, for example, during low rainfall but not drought conditions during most summer and fall seasons. These periods of relatively low flow may extend beyond the 60-d duration assumed for short-term PNECs by Caldwell et al. 21. As such, one can ask what fraction of U.S. waters would potentially be at risk from estrogens if the long-term PNECs are applicable to these extended periods of relatively low flow? Thirteen percent of river reaches across all watersheds included in PhATE have a low flow PEC greater than the long-term PNEC with a low of 4.8% (Columbia River) to a high of 70% (Raritan River) between watersheds (see Supplemental Data). These comparisons indicate a larger percentage of U.S. surface waters have PEC:PNEC ratios >1 than for the other comparisons presented above (i.e., low flow PECs to short-term PNECs and mean flow PECs to long-term PNECs; Table 2). However only 4.5 and 1.1% of river reaches have low flow PEC:long-term PNEC ratios of greater than 2 and 3, respectively, indicating the exceedances are not large. The comparison should also be viewed as a conservative upper bound estimate. The long-term PNECs developed by Caldwell et al. 21 were based on continuous exposure of test species for most of a single generation or multiple generations. Low summer and fall flows in most U.S. rivers can be interspersed with higher flows during which estrogen concentrations decrease and are, thus, not continuous. The effect of decreasing the exposure concentration for a portion of the life cycle is unknown. Additionally, few fish species go through their entire life cycle during the late summer and fall when flows may be lowest. Given the apparent need for continuous exposures over multiple life stages and multiple generations to observe effects above the long-term PNEC, and the absence of effects at such concentrations when only a portion of the life cycle is exposed (Caldwell et al. 21), it is likely that the comparison of critical low flow PECs to long-term PNECs overestimates the percentage of surface waters in which effects may occur.

The apparent low risk for the U.S. watersheds appears very different from a similar study carried out in the United Kingdom where about 30% of all segments were categorized as at risk based on a mean concentration 19, 43. It should be noted that the UK study used a very precautionary (lower than this U.S. study) 1 ng/L E2-eq PNEC. However, if the lower UK PNEC 19 is used to evaluate the potential risk associated with E2-eq in U.S. surface waters, the exceedance rises to approximately 1 and 11% of segments being at risk for mean and critical low flow, respectively. Thus, even when the lower, more precautionary, UK PNEC is applied to the U.S. watersheds, there still appears to be little widespread risk predicted for wildlife in the U.S. river systems compared to that in the United Kingdom.

Why is there a 22- to 64-fold lower risk in the U.S. mean flow situation compared to the UK mean concentration? The predicted PECs will depend on the per capita loadings of the individual steroid estrogens to the POTWs, the removal percentages assumed for different treatment types 39, the dilution available in the receiving segments, and the assumed rate of degradation of the compounds as they are transported along the river network. When the two modeling approaches were compared, it was found that there were indeed small differences in excretion rates and removal efficiencies used in the two risk assessments. Using the UK modeling approach 19 gives rise to an effluent approximately 1.6 to 2.2 times estrogenically stronger than used in this present U.S. study. Clearly this, in itself, does not explain the 22- to 64-fold lower risk in the U.S. compared to the UK situation. The in-stream degradation rates used for the U.S. and UK assessments were found to be very similar. The final factor affecting the PECs is the total dilution available for the effluent from the POTW within the watersheds, which is expected to be higher in the United States and is related to population density. The average population density of the modeled UK watersheds is about 10 times greater than the U.S. watersheds (241 vs. 26 people per km217, 43). Assuming POTW effluent flow is directly related to population density and river flow is independent of population density, this suggests an approximately 10-fold greater dilution of POTW effluent in the United States than the United Kingdom.

Relationship between E2-eq PECs and intersex incidence

When all watersheds are combined, no relationship between E2-eq PECs and either the incidence of intersex fish (Fig. 4) or the plasma concentration of VTG (Supplemental Data) reported in the recent study by Hinck et al. 36 could be found. When PECs and the incidence of intersex were compared within an individual watershed, no relationship between the two was apparent (Supplemental Data). The highest PECs were predicted in reaches of the Lower Colorado Basin, yet no incidence of intersex was reported in those reaches. These comparisons suggest that releases of steroid estrogens from POTWs were not the primary cause of the occurrence of intersex fish in U.S. surface waters reported by Hinck et al. 36. Blazer et al. 34 also found that intersex in fish occurred in areas not affected by POTWs, with livestock operations being a possible alternative source of endocrine disrupting compounds.

Margin of safety in U.S. watersheds

As discussed previously, the precision or accuracy needed in a model depends to some extent on how close the predictions are to levels thought to pose a concern. The MOS for every segment included in PhATE was plotted for mean and low flow conditions. This exercise revealed that approximately 92% of the river segments of these 12 U.S. watersheds enjoy a margin of safety greater than 10, and of these 59% enjoy a margin of safety greater than 100 (Fig. 5). Thus, even if the overall estrogenic potential of a POTW effluent was three times greater than estimated for the steroid estrogens alone, there would be little increase in the number of river reaches with PECs greater than PNECs. Such a threefold increase in estrogenic potential is not expected for most POTWs given that numerous studies have shown that, with few exceptions, steroidal estrogens contribute most of the estrogenic activity downstream of POTW effluents 44. The studies included in that summary confirm that steroid estrogens are the dominant source of measured estrogenicity in POTW effluents, and consideration of the contribution of other sources of estrogenicity will not change the conclusions of the margin of safety evaluation. Alternatively, even if there was an additional 300% of the estrogenic activity from sources other than humans in waters that receive POTW discharges, adverse reproductive effects in aquatic biota would still be predicted to occur in fewer than 3% of stream segments and induction of biomarkers of exposure to estrogens would be measurable in fewer than 10% of segments.

It has been suggested that although concentrations of estrogens and other endocrine disrupting compounds in the field may often not be high enough to directly adversely affect reproduction, they may nevertheless sensitize organisms to other stressors. Laboratory evidence, to date, has not supported this idea. For example, co-exposure to another class of chemicals (surfactants), or to low oxygen concentrations in the water, did not alter the sensitivity of fish to a mixture of estrogenic chemicals 45, 46. Although it is obviously theoretically possible that co-occurring stressors could lead to greater than anticipated effects, this idea awaits confirmation by well-designed, well-conducted studies before it can be accepted. The large MOS estimated by our analysis for steroid estrogen exposures mitigate the potential for adverse effects from such interactions.

It remains possible that a few species may be more sensitive than those tested so far and that a lower PNEC would be required to protect against this possibility. The MOS estimated for U.S. watersheds indicate that even if the long-term PNECs for all steroid estrogens were decreased by 10-fold (i.e., to 0.6, 0.2, 6, and 0.01 ng/L for E1, E2, E3, and EE2, respectively), the vast majority of rivers segments in the United States (about 94%) would still have a MOS of 1 or greater; the E2-eq PEC would be less than the PNECs, indicating no adverse effects would be expected in the vast majority of segments.

CONCLUSION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSION
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

Whether a realistic PNEC or a very precautionary PNEC is used, this present modeling study indicates that no widespread risk of endocrine disruption exists to wild fish in the United States from human-derived steroid estrogens. The principal reason for this finding is the relatively high dilution enjoyed by the United States compared to many other (particularly European) developed nations. This does not mean that no endocrine disruption due to human-derived estrogens is predicted, rather that such occurrences will be rare compared to the large area covered by the U.S. water systems and focused in headwater streams with POTWs in arid portions of the country, as well as in and just downstream of urban centers. Of course, there are other sources of steroid estrogens (agriculture) and there are xenobiotic endocrine disrupters from agriculture or industry still to consider. This apparent low risk of U.S. surface waters to endocrine disruption, due primarily to relatively large amounts of dilution, suggests that other adverse environmental effects due to pharmaceuticals and personal care products emanating from the human population are also likely to be less than in many other developed nations. It is hoped that the outputs of GIS water quality models such as PhATE can be linked with chemical and biological field programs to help test these conclusions in the future.

Acknowledgements

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSION
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

The authors thank J. Arris of ARCADIS for keeping up with and incorporating numerous revisions to the manuscript. P. Anderson and D. Pfeiffer thank PhRMA for supporting this research. A. Johnson and R. Williams thank NERC through the CEH science budget for supporting their contribution.

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  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSION
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information
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Supporting Information

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSION
  8. SUPPLEMENTAL DATA
  9. Acknowledgements
  10. REFERENCES
  11. Supporting Information

Additional Supporting Information may be found in the online version of this article.

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