Development of a Management Plan for Lake Attitash, Amesbury and Merrimac, Massachusetts
Prepared by Water Resource Services, Inc.
May 2016
Contents
Introduction and Background ...... 1 Methods and Approach ...... 3 Review of Nutrient Budgets...... 6 Oxygen Demand Assessment ...... 13 Internal Load Assessment ...... 18 Modeling of Watershed and Lake ...... 21 Target Phosphorus Concentration ...... 25 Management Options Review ...... 26 Oxygenation Potential ...... 43 Phosphorus Inactivation Potential...... 45 Recommendations ...... 51 References ...... 52
Figures
Figure 1. Lake Attitash and General Vicinity ...... 2 Figure 2. Lake Attitash Watershed ...... 2 Figure 3. Lake Attitash Sediment Sampling Locations ...... 5 Figure 4. Lake Attitash Secchi Transparency from 1977 to 1998 (from CDM 1999) ...... 7 Figure 5. Lake Attitash Secchi Transparency Summary (from USEPA 2014) ...... 7 Figure 6. Lake Attitash Land Use Map (from CDM 1999) ...... 8 Figure 7. Hypolimnetic Phosphorus Concentration Change (ug/L) in Lake Attitash in 2011 ..... 11 Figure 8. Epilimnetic Phosphorus Concentration Change (ug/L) in Lake Attitash in 2011 ...... 11 Figure 9. Dissolved Oxygen Profiles from 1998 (From CDM 1999) ...... 14 Figure 10. Extent of Anoxia from 2011 (From USEPA 2014) ...... 15 Figure 11. Development of Anoxia in Lake Attitash in 2015...... 17 Figure 12. Development of Anoxia in Lake Attitash in 2015...... 17 Figure 13. Phosphorus in Lake Attitash in 2011 (From USEPA 2014)...... 19 Figure 14. Land Based (left) and Underwater (right) Elements of a Diffused Oxygen System ... 44 Figure 15. Schematic of a Sidestream Supersaturation Oxygenation System ...... 44 Figure 16. Application of Aluminum...... 46 Figure 17. Aluminum Assay Results for Lake Attitash Sediment ...... 48
Tables
Table 1. Phosphorus Budget from 2011 Data (from USEPA 2014) ...... 10 Table 2. Calculation of Oxygen Demand ...... 18 Table 3. Calculation of Phosphorus Release from Sediment in Lake Attitash ...... 19 Table 4. Lake Attitash Sediment Features in 2015 ...... 20 Table 5. Land Use Data for the Watershed of Lake Attitash ...... 22 Table 6. Loading Features for the Three Drainage Areas of Lake Attitash ...... 22 Table 7. Total Phosphorus (ug/L) Based on 2001-2011 Data for the Lake Attitash System ...... 22 Table 8. Loading Summary for Current Conditions in Lake Attitash ...... 23 Table 9. Loading Comparison for Lake Attitash Management Scenarios ...... 23 Table 10. Options for Control of Algae and Floating Plants (Adapted from Wagner 2001) ...... 27
Introduction and Background Lake Attitash covers 360 acres (145 hectares) in Amesbury and Merrimac, Massachusetts, not far from the New Hampshire border (Figure 1). The lake has a maximum depth of 32 feet (9 meters) and an average depth of just under 12 feet (3.6 meters). The watershed covers almost 2500 acres (just over 1000 hectares) and extends into New Hampshire. The main tributary is the Back River, entering from the west-northwest, with a much smaller secondary tributary entering from the southwest and additional direct drainage around the lake (Figure 2). The lake is used primarily for recreation, but is also a secondary water supply for the Town of Amesbury, and there are nearby wellfields. It has a public boat ramp and is heavily used by shoreline residents as well as day trippers.
An excellent historical timeline has been provided by the USEPA (2014). Historically, Lake Attitash was used for water power for downstream mills, with its level raised by 3 feet in 1712 and water level issues continuing into the 1970s. Lake Attitash was a popular destination going back 100 years, and had many shoreline summer cottages with wells and on-site waste disposal systems. There were camps, dance halls, private beaches, and a host of supporting services. Popularity for day trips and development increased in the 1960s with the completion of Route 495. Until the Wetlands Protection Act was passed, considerable wetland area was impacted.
The hydropower rights were sold to the Town of Amesbury in the mid-1960s and the lake became a back-up water supply, with the potential to release flows that reached the town intake downstream on the Powwow River. Conversion of summer places to year round dwellings progressed through the 1970s, with additional development in the 1980s, and public water and sewer were provided to most residences by the early 1990s. There has been some development since then, mostly as camps or other larger properties were sold and redeveloped, but additional building space is limited.
Agricultural use has been substantial in the watershed, mainly in the drainage area of the Back River. The installation of tile drainage at the Sargent farm in the 1960s is believed to have greatly increased loading to the river, especially after the farm started accepting waste for composting. The addition of gelatin waste from Kraft Foods by permit in 2005 and possibly as early as 1989 less formallywas viewed as particularly problematic and was ceased in 2010. A number of farm drainage improvements have been made since the mid-1990s.
There is single report of algae problems from 1944, but rooted plants were more of a problem through most of the twentieth century, suggesting limited algae blooms. The lake was considered to be moderately fertile in the 1970s, with reported blooms only late in that decade. Conditions have deteriorated since then, and Lake Attitash has experienced algae blooms for at least three decades, but with increasing frequency, severity and dominance by cyanobacteria. Algae blooms over the last decade have included the cyanobacteria Dolichospermum (formerly Anabaena), Aphanizomenon and Microcystis, with less abundant Aphanocapsa, Woronichinia, Pseudanabaena, Oscillatoria (probably actually Planktothrix), Gomphosphaeria, Anabaenopsis, Cylindrospermum and Synecchococcus or Cyanobium. Additionally, chorococcalean green algae, euglenoids and some dinoflagellates have been abundant at times. Summer algae biomass has typically been in the range of 3-8 mg/L, an elevated range. Taste and odor events have occurred, and toxicity (microcystin) has been detected.
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Figure 1. Lake Attitash and General Vicinity
Figure 2. Lake Attitash Watershed
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Rooted plants include invasive Eurasian water milfoil and water chestnut as well as a range of native species, and nuisance densities are achieved in shallow water in many summers. However, the low clarity limits the extent of rooted plant coverage and density in deeper water. Water clarity is typically between 3 and 7 feet (1-2 meters) for most of the summer.
Lake Attitash has been the subject of multiple investigations over several decades. Massachusetts state agencies studied the lake in the 1970s, and CDM conducted a major evaluation in the late 1990s. The USEPA performed a substantial study and review of available data in 2011, with a report in 2014. Additional data have been generated by the Lake Attitash Association (LAA) through summer monitoring and by the University of New Hampshire as part of spring and fall class activities and independent thesis research. The 1999 CDM report suggested management needs and options, and the LAA has followed up on many needs with the aid of Amesbury, Merrimac and various state agencies. Grants have been received and both watershed and in-lake management efforts have been conducted. The 2014 USEPA report assesses the results of more than a decade of effort. While some improvement has been noted, conditions remain degraded from the perspective of summer recreational use and water supply.
This investigation was conducted to evaluate the need for in-lake action to remediate excessive phosphorus and algae. It includes:
1. Review of past nutrient budgets and further consideration of internal loading. 2. Oxygen demand estimation to determine the potential for reduced internal loading through oxygenation. 3. Sediment assessment to evaluate available phosphorus and the potential for control through inactivation of surficial sediment. 4. Derivation of a target phosphorus concentration for Lake Attitash that would acceptably minimize algae blooms and evaluation of the potential for internal load control to reach that target. 5. Evaluation of all applicable techniques to reduce phosphorus concentrations in Lake Attitash.
Methods and Approach Review of past nutrient budgets and further consideration of internal loading involves mainly evaluation of the CDM (1999) and USEPA (2014) reports on Lake Attitash. Consideration of watershed management efforts and additional monitoring data is warranted, but these two reports constitute the best available assessments.
Oxygen demand estimation to determine the potential for reduced internal loading through oxygenation involves weekly to biweekly oxygen profiles between ice out and whenever anoxia occurs off the lake bottom, allowing calculation of oxygen demand. Continued oxygen assessment facilitates determination of the areal extent of anoxia and the zone of internal phosphorus loading. Oxygen demand cannot be properly estimated from profiles in which more than the bottom value is <2 mg/L, and is best derived from mid-spring data. The maximum areal extent of anoxia has been depicted in the USEPA (2014) report, but additional characterization of variability was desired.
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Sediment assessment to evaluate available phosphorus and the potential for control through inactivation of surficial sediment involves collection of sediment samples and testing for phosphorus bound to iron and subject to release under anoxia. Iron-bound phosphorus is viewed as the available phosphorus in relation to anoxic releases, as reactions in the absence of oxygen cause iron and phosphorus to dissociate and become soluble, moving into the overlying water if not already saturated with each. Surficial sediment samples were collected with an Ekman dredge from 6 locations (Figure 3). Samples representing the upper 10 cm of sediment were tested for percent solids, total phosphorus and iron-bound phosphorus by Northeast Laboratories of Berlin, CT. At one location (site 267 on Figure 3) a core sample was obtained with a gravity corer and additional 4 cm slices were assessed as for the upper 10 cm samples to determine vertical variation in available sediment phosphorus.
Additionally, aluminum assays were conducted by Northeast Laboratories on two individual samples and one composite sample to evaluate the reduction in iron-bound phosphorus with increasing aluminum application. Small amounts of sediment are treated with aluminum compounds and then retested for iron-bound phosphorus. This allows a projection of the level of reduction in internal loading that might be obtained by an inactivation treatment of surficial sediment and facilitates dose planning and cost estimation.
In order to evaluate the possible benefits of internal load control, the Lake Loading Response Model (LLRM, AECOM 2009) was applied. LLRM is a spreadsheet model that allows actual data to be used to set variable values and provide reality checks on calculated values at key points in the modeling process. The most recently available land use data and the 2011 water quality data from USEPA (2014) were applied, tempered with longer term average values from annual monitoring of Lake Attitash. Loading coefficients and attenuation factors were adjusted until the predicted stream input and in-lake values were close to actual values. Once the model was considered to represent current conditions in Lake Attitash, changes in external and internal loading were evaluated to determine the potential impact from possible management actions.
Derivation of a target phosphorus concentration for Lake Attitash that would acceptably minimize algae blooms is aided by LLRM. Varying phosphorus concentrations allows prediction of chlorophyll and Secchi transparency and consideration of acceptable conditions.
Evaluation of applicable techniques to reduce phosphorus concentrations in Lake Attitash focuses on oxygenation and phosphorus inactivation approaches, as these are the most practical measures for achieving substantial in-lake phosphorus concentration reductions and associated decline in algae abundance. Other techniques are also evaluated, including additional external loading efforts, hypolimnetic withdrawal and dredging. LLRM was applied to assess possible reductions in phosphorus and algae, and the logistics and costs of possible management efforts were considered.
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Figure 3. Lake Attitash Sediment Sampling Locations
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Review of Nutrient Budgets The CDM (1999) report is based on water quality data collected in 1994-1998, with consideration of USEPA data from 1977-1978. The pattern of water clarity over time (Figure 4) indicates that there were issues even in the 1970s but that conditions have worsened over time; further monitoring by the Lake Attitash Association (LAA) and the USEPA since 2000 have confirmed the degraded water clarity during summer (Figure 5). Conditions may not have worsened appreciably since the 1990s, but even with watershed management efforts, water clarity has not improved.
CDM estimated the loading of phosphorus to Lake Attitash by two approaches: multiplication of concentrations and flows for tributaries and other sources from limited data and modeling of loading from land use data and export coefficients. Actual values applied by CDM included very high nitrate nitrogen levels in the Back River in 1998, with occasional spikes in the Southwest tributary inlet but generally low levels in Lake Attitash. A similar pattern was noted for ammonium nitrogen, and activities on the Sargent Farm upstream on the Back River may have been responsible for elevated values. Erratic but generally high total phosphorus was observed in the Back River, with some extremely high levels near Sargent Farm. A few very high values resulted in a mean of 43 ug/L for surface water in Lake Attitash, but the median value was 22 ug/L. The CDM report also noted that there was a decline in phosphorus near the mouth of the Back River. The substantial sediment build up in that area suggests that much particulate phosphorus is settling out before it is available to algae in the water column of the lake.
The loading from multiplication of flows and concentrations suffered from limited data and a few very questionable phosphorus values that suggest lab error; values of this magnitude are rarely encountered, and while some very high values are possible in the Back River or storm drainage systems during storms, extreme values should have been rejected as outliers. At the least, median values should have been applied instead of the mean. As a result, the loading estimates based on this approach are extremely high and not consistent with observed in-lake conditions. USEPA (2014) also questioned the CDM load estimates from this direct approach.
The alternative approach, using literature values for export from various land uses and land areas in each use, has been applied in many studies and has been found to be a fairly reliable way of getting at least a rough estimate of loads. It is best when combined with actual data to adjust the export coefficients, but can provide a reasonable assessment of loading. Land use appears to have been from 1995 (Figure 6), and the resulting analysis suggests an external phosphorus load of 365 kg/yr, with 47% from the Back River, 40% from direct drainage, and 13% from the southwest inlet (unnamed tributary). The applied export coefficients are consistent with expectations for this and similar areas in New England. Assuming no other loads or attenuation of external loading upon entry to the lake, empirical equations for estimating load from in-lake concentration, depth and flushing rate suggest an average phosphorus concentration between 22 and 29 ug/L in the 1990s. This is more consistent with values from that decade that exclude extreme (outlier) data and results in predicted chlorophyll (12 ug/L) and Secchi transparency (1.8 m) values that are similar to those actually recorded.
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Figure 4. Lake Attitash Secchi Transparency from 1977 to 1998 (from CDM 1999)
Figure 5. Lake Attitash Secchi Transparency Summary (from USEPA 2014)
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Figure 6. Lake Attitash Land Use Map (from CDM 1999)
The CDM (1999) report noted anoxia in the hypolimnion (bottom water layer during summer stratification) and suggested that this condition was worse than in the 1970s. An increase in the areal extent of summer anoxia could be a major factor in increased internal phosphorus loading and the observed decline in water clarity. It is not clear that there was any major land use change between the 1970s and 1990s, but provision of sewers in the 1980s and early 1990s should have reduced loading and limited external influences on the lake. While the phosphorus ultimately came from the watershed, the relatively long detention time for Lake Attitash would allow much of that phosphorus to accumulate in bottom sediments, and loss of oxygen would facilitate internal recycling of phosphorus sufficient to support algae blooms without excessive external phosphorus inputs.
The CM (1999) report also noted that cyanobacteria were the dominant algae in the 1990s. Algae blooms were not reported for the period of data collection in 1977-1978, but were apparently observed at some point in between the two monitoring periods, although the types of algae are not given. Cyanobacteria tend to be favored by higher internal loading, as the N:P ratio of associated loads are low; nitrogen limitation is avoided by many cyanobacteria by virtue of their ability to use dissolved nitrogen gas, something most algae cannot do. The sewering completed
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in the early 1990s should have raised the N:P ratio, as nitrogen is much more mobile than phosphorus through soil, so again internal loading is implicated as an increasing influence on Lake Attitash as the watershed influences waned. The CDM report noted the potential importance of internal phosphorus loading on Lake Attitash and suggested additional investigation, but most of the management recommendations focused on watershed management to reduce external loading.
CDM did not undertake a complete nitrogen assessment, but nitrate and ammonium values in Lake Attitash surface water in 1997-1998 were generally low. High values were sometimes observed in tributaries, and residual inputs from ground water influenced by past on-site wastewater systems may have been elevated at that time, but it appears that nitrogen may have been a limiting factor for many algae by the late 1990s. This is consistent with cyanobacteria blooms when phosphorus levels are adequate to support them, which the available data suggests was the case. However, other than those few outlier data points, phosphorus was not extremely high, and the role of internal loading could be very important in supporting those blooms.
Hillary Snook of Region I of the USEPA undertook a water quality assessment in 2011 and summarized those results and many past efforts in a 2014 report. Considerable direct measurements were made on surface water, including separation of dry and wet weather inputs. Atmospheric and ground water phosphorus inputs were estimated from calculated flows and literature values for average content, and internal loading was estimated as the difference between surface and bottom phosphorus levels times the volume of the hypolimnion. The resulting phosphorus loading breakdown (Table 1) suggests that external phosphorus loading was on the order of 554 kg/yr (1219 lbs/yr), but contains some estimates that are likely high.
Inflow from direct precipitation is listed as 1.77 cfs, equivalent to about 43 inches per year, an appropriate value for the Lake Attitash area. However, precipitation in this region rarely has an average phosphorus content >20 ug/L (0.020 mg/L), and we usually use an average value of 17 ug/L based on many past measurements. This suggests that the atmospheric load should be 27 to 30 kg/yr, substantially lower than the 47.5 kg/yr (105 lbs/yr) value in Table 1.
Ground water could have a concentration near 20 ug/L, as suggested in Table 1, but the rate of inflow seems very high. There is no way to be sure without direct measurement, but assuming that all areas without substantial muck cover (roughly the 4-4.5 m contour) contribute seepage, the total interface area would be about 700,000 m2. Assuming an inflow of 10 L/m2/day (an average value for areas with sandy soils), that would translate into a lakewide ground water input of 2.83 cfs, much lower than the 6.11 cfs given in Table 1. And the phosphorus concentration in ground water should be closer to 10 ug/L in the absence of on-site waste water disposal, so the ground water phosphorus load is likely to be on the order of 25 kg/yr, much lower than the 109 kg/yr (241 lbs/yr) estimated in Table 1. Collecting some direct data for ground water inputs, both quantity and quality, is recommended, but a lower estimate of direct ground water phosphorus loading is warranted until appropriate data are obtained.
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Table 1. Phosphorus Budget from 2011 Data (from USEPA 2014)
The surface water inflows and phosphorus content are based on actual data, and should be reasonable estimates for 2011 based on twice monthly sampling. However, the multiplication of mean values introduces error; the summation of individual loads based on the product of concentration times flow can be very different than the mean of all concentrations times the mean of all flows. Further, 2011 was a wetter than average year, particularly in June, August and October, and would be expected to have higher flows and greater non-point source loading. Frequent and accurate measurements are needed for an accurate loading analysis when non-point source loading is dominant, and while the USEPA effort was quite impressive, it may not be adequate to account for associated variability.
In particular, the average outflow, which appears to have been carefully derived, translates into an average water yield of >4 cfs per square mile of watershed, while average values for this region are 1.5 to 2.0 cfs per square mile. The combination of measurement dates and wetter weather may be partly responsible. Also, the lake level appears to have been almost a foot lower at the end of the USEPA study than at the start. This alone would equate to 0.5 cfs of “extra” flow at the outlet. Longer term average loads should be less than those derived for 2011 from actual data. The summation of the surface water inflows is 397 kg/yr. The corresponding loading from the CDM actual data was much higher but deemed erroneous. The corresponding 1999 loading from the CDM land use model was 365 kg/yr, lower but not very different, but also representing a period before some important watershed best management practices were instituted.
The estimation of internal load can be performed by multiple methods, depending on the available data. The USEPA (2014) effort chose to estimate internal load as the concentration difference between surface and bottom samples times the volume of the hypolimnion, deriving a value of 53.3 kg/yr (117.3 lb/yr in Table 1). This will be an underestimate of the phosphorus released from surficial sediment, as there is continual flux going on. The change in concentration over time in the hypolimnion (Figure 7) reveals this flux.
During the period of stratification and bottom anoxia at the measurement station, which is roughly June through August, the concentration increases in early June, decreases with an apparent mixing event in early July, and oscillates several times before stratification ends in
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Date to Date Concentration Difference at Bottom 40
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9/7/2011 4/6/2011 5/4/2011 6/1/2011
4/20/2011 5/18/2011 6/15/2011 6/29/2011 7/13/2011 7/27/2011 8/10/2011 8/24/2011 9/21/2011 10/5/2011 11/2/2011
11/16/2011 11/30/2011 10/19/2011
Figure 7. Hypolimnetic Phosphorus Concentration Change (ug/L) in Lake Attitash in 2011
Date to Date Concentration Difference at Top 40
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4/20/2011 5/18/2011 6/15/2011 6/29/2011 7/13/2011 7/27/2011 8/10/2011 8/24/2011 9/21/2011 10/5/2011 11/2/2011
11/16/2011 11/30/2011 10/19/2011
Figure 8. Epilimnetic Phosphorus Concentration Change (ug/L) in Lake Attitash in 2011
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September and a major decline is observed. Much of the decrease in hypolimnetic phosphorus concentration in June through August will correspond with increases in epilimnetic phosphorus level, but with the epilimnion having more than 5 times the volume of the hypolimnion, the change is within the margin of error for laboratory measurement and is not visually evident (Figure 8). When stratification ends, not all of the hypolimnetic phosphorus is mixed with the epilimnion; adding oxygen will cause a substantial amount of phosphorus precipitation, so care must be taken with data interpretation.
There are other ways to use the available data to estimate phosphorus release from sediment, and additional data have been collected as part of the WRS effort to enhance internal load evaluation, but these will be handled later in the section on internal load assessment.
The USEPA report discusses major influences on nutrient budgets, including both phosphorus and nitrogen, although no loading analysis is provided for nitrogen. The history of activities in the watershed and lake has major bearing on current conditions. The trenching that has short circuited wetlands to alleviate flooding or open land for development prior to passage of the Wetlands Protection Act has limited the nutrient removal function of those wetlands. Similar activities on agricultural lands, notably the installation of tile drainage, has reduced attenuation and increased loads to tributaries, especially the Back River. Storm water drainage systems associated with residential development protected property from storm damage but routed contaminated water more quickly to the lake, increasing loads. Both hydrology and nutrient loading have been changed in ways that negatively impact Lake Attitash over the last 100 years. However, nearly all activities that set up direct negative consequences to the lake were taken before awareness and regulation of the impacts of those activities were completely understood.
Implementation of best management practices (BMPs) in the watershed has been ongoing since the 1980s. Certainly the provision of sanitary sewers was a major development, and more recent storm water management improvements have been significant. Certainly more could be done, but considerable and enlightened effort has been expended locally and supported at the state and federal level through grants and in-kind services. The USEPA report laments the lack of follow up monitoring and maintenance for many BMPs, a common problem in many watersheds and the greatest limitation on the success of the Section 319 program under the Clean Water Act. Still, the actions taken over the last three decades should have improved water quality to some degree, but improvement in the lake has not been documented. To the contrary, cyanobacteria blooms have increased in frequency and possibly severity. The increased extent of anoxia and potential for internal phosphorus loading are cited as likely factors, and WRS concurs.
While the loading analyses provided by CDM and the USEPA are limited by available data, consideration of the fundamental sources and general order of magnitude is consistent between them. It is clear that watershed loading increased through at least the 1970s, and possibly until early in the 1990s, and that actions since then should represent improvements to water quality. However, the level of loading from each analysis, as adjusted by this review, suggest fairly similar loading and similar in-lake levels of at least phosphorus, both of which are excessive. The one open question is the level of internal phosphorus loading, which does appear to have increased between the 1990s and today. We will further investigate external and internal loading in the section of this report on watershed and lake modeling.
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Oxygen Demand Assessment Loss of oxygen in deep water has major impacts on lake ecology and uses. Low oxygen restricts habitat for fish and invertebrates, and complete loss of oxygen (anoxia) fosters chemical and biochemical reactions that either release certain contaminants from the surficial sediment or prevent processing of contaminants that enter the deep water from upper waters. Decomposition slows, ammonium cannot be converted to nitrate, and phosphorus can be released from sediment as it dissociates from iron compounds which often represent a substantial fraction of the total phosphorus in those sediments. Anoxia is not uncommon in lakes that stratify during summer, as the movement of oxygen downward across the boundary between upper and lower water layers (the thermocline) is very slow and decomposition and related bacterial metabolism consumes the available oxygen faster than it can be replaced. However, elevated loading of organic matter and nutrients from the watershed accelerates this process, causing faster and more severe anoxia over a greater area than might otherwise occur.
There is a general perception that anoxia has increased in Lake Attitash over time, and while this is a logical progression that may well be true, the supporting data are very limited. The 1977- 1978 study only measured oxygen to 6 m (20 feet) and found no low oxygen. While more recent measurements have suggested anoxia at shallower depths, it is possible that oxygen at the bottom in >9 m (30 feet) of water may have been low even in the 1970s. The 1994 oxygen levels were <1 mg/L from 6-10 m (20-33 feet) from mid-July through at least mid-August, but no true anoxia was reported. However, the technology for oxygen measurement by probes was not as advanced at that time and values <1 mg/L were often assumed to represent anoxia.
In 1998 oxygen was very low below 6 m (20 feet) in early August, below 5 m (16.5 feet) in mid- August, and below 6 m again in late August, with low oxygen below 8 m (26 feet) by early September (Figure 9). Thermal stratification put the thermocline at about 4.5 m (15 feet), with breakdown in early September. While low values were <1 mg/L and not measured as 0 mg/L, the probe technology issue is again raised and anoxia may well have existed. Certainly the pattern suggests complete loss of oxygen by sometime in July with the anoxic interface creeping upward through mid-August and then moving downward later in August and dissipating in September.
The 2011 data from the USEPA indicate that anoxia occurred below a depth of 7.5 m by mid- June, below 5.5 m by the end of June, below 4.5 m in July and August (Figure 10). Anoxia was also detected below a depth of 3.5 m in February; winter anoxia does not appear to have been evaluated previously. Even if we assume that anoxia was present in earlier years, the duration and extent of anoxia does appear to have expanded over the last few decades. This increases the potential for internal loading, especially when the anoxic interface approaches the thermocline. Under those conditions, algae can grow near the thermocline, getting enough light from above and ample phosphorus from below. The surface water phosphorus concentrations are marginal for severe bloom support; bloom formation is likely to involve growth and luxury uptake of phosphorus from either the thermocline area or directly from sediment in areas where some light penetrates.
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Figure 9. Dissolved Oxygen Profiles from 1998 (From CDM 1999)
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Figure 10. Extent of Anoxia from 2011 (From USEPA 2014)
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Measurement of oxygen in 2015 (Figure 11) from mid-April through late June illustrate the development of anoxia in Lake Attitash. Oxygen levels exhibit just a slight decline near the bottom through early May, with a sharper decline but no anoxia in mid-May and the first evidence of anoxia in late May. By late June anoxia has expanded from areas deeper than 8 m in late May to areas deeper than 6 m. This progression (and that of later summer profiles not shown for simplicity here) is similar to that observed in 2011. The progression does not completely track the thermal profiles (Figure 12), which exhibit the start of stratification in early May, show substantial stratification by mid-May, but demonstrate some mixing and stratification breakdown in late May through late June. Stronger stratification eventually occurs in 2015, but the June mixing may well have allowed phosphorus released from an anoxic zone covering between 6 and 22 acres during June to be mixed into upper waters where it supports algae growth. Such mixing events may occur at other times later in summer as well, increasing phosphorus availability with limited predictability.
The data from spring of 2015 were collected at intervals intended to support calculation of oxygen demand (Table 2). Spring data are best for this purpose, as oxygen uptake is hindered at low oxygen levels; oxygen demand can be most easily expressed (and therefore more accurately measured) when all values are higher than about 2 mg/L. Below that level, the uptake kinetics are altered and underestimation of actual demand is likely. Data from 3 profiles meet the desired criteria for optimal oxygen demand assessment: 4/28/15, 5/5/15 and 5/14/15. Some vertical oxygen flux may introduce error to these calculations, but using data from the water column below the point of eventual stratification usually provides valid results. Consequently, the difference in oxygen level between corresponding values at each 1 m increment from 4 to 9 m of depth was summed to represent the loss of oxygen over the period of time between each pair of measurements. These differences are adjusted to account for loss of oxygen due to temperature increase. Rising temperature occurs over the targeted period and higher temperatures create lower oxygen saturation levels; not correcting for that factor will cause an overestimate of oxygen demand.
The results are expressed as oxygen loss in g/m2/day, with values over about 0.5 g/m2/day usually causing eventual anoxia over the period of stratification and values >4 g/m2/day causing rapid loss of oxygen. The values for Lake Attitash based on the spring 2015 data are just over 1 g/m2/day, a substantial but not overwhelming oxygen demand. This seems consistent with observed oxygen profiles; it takes until late June for a substantial area and volume of the lake to go anoxic. While the spacing of profiles in 2011 was not ideal, 3 estimates were derived from those data, using dates of 5/3/11, 5/24/11, 6/2/11 and 6/14/11. Values were 0.5, 0.5 and 1.0 g/m2/day. Oxygen demand is therefore great enough to cause anoxia in Lake Attitash, but not so high as to cause rapid anoxia as stratification sets up. The relatively quick return to oxygenated conditions with late summer mixing is also consistent with projected moderate oxygen demand.
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Oxygen Profile Progression in Lake Attitash in 2015 Dissolved Oxygen (mg/L) 0.0 2.0 4.0 6.0 8.0 10.0 12.0 0 1 4/16/2015 2 4/28/2015 3 5/5/2015 4 5/14/2015
5 5/31/2015 Depth(m) 6 6/27/2015 7 8 9
Figure 11. Development of Anoxia in Lake Attitash in 2015
Temperature Profile Progression in Lake Attitash in 2015 Temperature ( deg C) 0.0 5.0 10.0 15.0 20.0 25.0 30.0 0 1 4/16/2015 2 4/28/2015 3 5/5/2015 4 5/14/2015
5 5/31/2015 Depth(m) 6 6/27/2015 7 8 9
Figure 12. Development of Anoxia in Lake Attitash in 2015
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Table 2. Calculation of Oxygen Demand
DO mg/L DO mg/L Depth (m) 4/28/2015 5/5/2015 Diff Temp Adj Depth (m) 5/5/2015 5/14/2015 Diff Temp Adj 1 9.8 8.8 1 8.8 7.7 2 9.9 8.9 2 8.9 7.7 3 9.8 8.8 3 8.8 7.7 4 9.8 9.3 0.6 0.3 4 9.3 7.6 1.6 0.7 5 9.8 8.7 1.1 0.8 5 8.7 6.7 2.0 1.1 6 9.8 8.2 1.6 1.3 6 8.2 5.9 2.3 1.3 7 9.7 7.9 1.8 1.5 7 7.9 5.3 2.6 1.7 8 9.7 7.5 2.2 1.9 8 7.5 4.2 3.3 2.4 9 9.4 7.3 2.1 1.8 9 7.3 3.3 4.0 3.1 Sum of diffs 9.33 7.53 Sum of diffs 15.86 10.16 Vertical interval (m) 1 1 Vertical interval (m) 1 1 Conv to g total 9.33 7.53 Conv to g total 15.86 10.16 Days between measures 7 7 Days between measures 9 9 HOD (g/m2/d) 1.33 1.08 HOD (g/m2/d) 1.76 1.13
Internal Load Assessment The 2011 data for phosphorus concentration in the two layers of Lake Attitash during summer stratification (Figure 13) are useful for estimating potential internal loading. In the USEPA report, the difference in concentration between the two layers over the course of stratification was used in conjunction with the hypolimnetic volume to estimate phosphorus accumulation in the hypolimnion as an estimate of internal loading. As described previously, however, this ignores the flux of phosphorus out of (or possibly into) the hypolimnion over this period, and is likely to underestimate actual release from surficial sediment.
The data are better used to evaluate release rates, much as the oxygen demand was estimated, using shorter periods of time within the period of stratification during which anoxia is present, mixing appears minimal and changes in concentration are most likely to represent actual releases from sediment. Based on the concentration change pattern in Figure 7, there are 3 periods where phosphorus accumulation appears suitable for estimating release: June 2-14, July 12-26, and August 9-23 in 2011. Multiplying the increase in hypolimnetic concentration during each time period by the corresponding volume of the hypolimnion and dividing by area, estimated release rates can be derived (Table 3).
The value for phosphorus release from sediment was 1.2 mg/m2/day for the June period, which had a relatively small hypolimnetic thickness and volume and a lesser change in concentration. Recall that stratification is still setting up in June and anoxia has only begun. While oxygen concentrations cannot be <0 mg/L, the redox potential (which translates into unfulfilled demand for oxygen) can go negative and continue to decline as anoxia persists. As redox potential goes
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Figure 13. Phosphorus in Lake Attitash in 2011 (From USEPA 2014)
Table 3. Calculation of Phosphorus Release from Sediment in Lake Attitash
Days Hypolimnion Increase in Date Range Elapsed Avg Depth Conc Release Rate (m) (ug/L) (mg/m2/day) 6/2-14/11 12 0.77 19.1 1.2 7/12-26/11 14 3.15 33.9 7.6 8/9-23/11 14 3.15 31.8 7.2 even more negative, reactions that release phosphorus accelerate and result in larger releases per unit area and time. For the July and August periods, with anoxia over the maximum depth range and area and continued redox depression, the calculated phosphorus release increases to 7.6 and 7.2 mg/m2/day, respectively. The range of values typically encountered in anoxic zones where iron-bound phosphorus in the sediment is common is about 2 to 20 mg/m2/day, with most values falling between 6 and 12 mg/m2/day. The July and August values are very much in line with expectations from the literature.
At an average of 7.4 mg/m2/day for 60 days over July and August, with an area of 526,000 m2 (130 acres, the area below a depth of 4.5 m), the total release would be 233.5 kg. Release may be interrupted by mixing events that bring some oxygen into deeper water and raise the redox potential, and not all of the released phosphorus will necessarily make it to the upper waters. On the other hand, this analysis does not consider any winter release that may indeed occur in light
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of observed anoxia in the one winter profile from 2011. As a rough estimate, about half of this estimated total release might be expected as an effective internal load, or about 117 kg/yr. This is unlikely to be an overestimate, but is more than twice the value suggested by the USEPA (2014) report.
Another approach to estimating internal phosphorus loading involves actual measurement of iron-bound phosphorus in the surficial sediment and estimation of what portion might be released over a given time period, usually a summer stratification period. There is substantial variability in sediment features (Table 4), and the iron-bound phosphorus measurements tend to be inversely related to solids content. The core sample exhibits a steady increase in solids and decrease in iron-bound phosphorus with increasing depth into the sediment. Depending on which values are used to create an average and any area based weighting, mean iron-bound phosphorus is on the order of 350 to 650 mg/kg. Using 406 mg/kg and an average solids content of 11.7%, the estimated iron-bound phosphorus mass in the upper 10 cm of sediment over the area subject to anoxia during summer is about 4087 kg. Using 510 mg/kg, the phosphorus mass is 5135 kg. These are very large reserves of potentially available phosphorus.
It is possible that as little as the upper 4 cm interact with the overlying water, and it would be unusual for more than about 10% of the iron-bound phosphorus to actually be released over the course of one stratification period, so the actual mass of released phosphorus might be no more than 163 to 205 kg/yr. This range falls in between the 233.5 kg estimate based on average release rate for 60 days and the 117 kg estimate of what portion of that release might actually reach the upper waters. In all cases this is a large potential internal load that might support algae blooms independent of external loads.
One other interesting aspect of the Lake Attitash sediment comes from the USEPA effort, which determined that iron levels in surficial sediment were between 26,000 and 44,000 mg/kg, while aluminum levels were between 11,000 and 17,000 mg/kg. Recent work in Maine (Norton and Amirbahman, unpublished) suggests that the aluminum levels have to be higher than the iron levels to avoid significant phosphorus releases under anoxic conditions, possibly as high as threefold. Lake Attitash would be considered an iron dominated system, susceptible to high internal loading when the hypolimnion goes anoxic.
Table 4. Lake Attitash Sediment Features in 2015
Total Iron-Bound % Total Phosphorus Phosphorus Sample ID # Solids (mg/kg) (mg/kg) 267SED 0-4 11.0 2390 1114 267SED 4-8 13.0 838 430 267SED 8-12 16.0 1250 220 267SED 12-16 21.0 1897 209 269SED 13.0 1460 348 270SED 11.0 1250 440 271SED 28.0 604 184 272SED 9.5 1200 1073 273SED 11.0 1760 429
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Modeling of Watershed and Lake The LLRM model was applied, which is a simple spreadsheet model that relates land use and related features to nutrient loading and uses simple empirical equations long accepted for this purpose to estimate resulting in-lake conditions. Land use data from 2010 were applied and export coefficients for water, phosphorus and nitrogen representative of similar land uses in the region for which actual export values could be calculated from downstream data. Land use was somewhat different than for the 1998 modeling effort by CDM (Table 5), which was based on 1995 land use data. Differences suggest less developed land in 2010 which is very unlikely; the 2010 land use data from Massachusetts and New Hampshire state GIS databases probably categorized some residential land as forest based on tree cover. This may be appropriate, or may cause slight underestimation of loading from that land. Agricultural land decreased slightly between 1995 and 2010, which is believable, and forest is listed as an increase; the underestimation of residential land is probably embodied in this likely overestimation.
Model calibration did not require adjustment of any export coefficients or attenuation factors outside the rational range for this area to get reasonable agreement between model predictions and known values for inlets. Mixed use watersheds tend to have average export levels on the order of 0.2 to 0.4 kg/ha/yr, and values for the current three drainage basins ranged from 0.22 to 0.36 (Table 6). Export coefficients for specific land uses were similar to those applied in the CDM modeling effort. Water load and phosphorus and nitrogen concentrations were <7% different between predictions and “reality check” values based on data.
However, the resulting loads for the three defined drainage areas were different than those derived by CDM, with the Back River accounting for almost 64% of the current load compared to 47% in the CDM model result. The direct drainage area contributed 27% of the external surface water load, compared to 40% in the CDM model, while the southwest tributary provided about 9%, compared to 13% in the CDM analysis.
Atmospheric loading was estimated as the volume of direct precipitation times the average phosphorus and nitrogen concentrations in rainfall in this region from other studies. Internal loading was based on the 7.4 mg/m2/day value derived from 2011 data for July and August, with a 60 day period and 130 acre area of contribution. Waterfowl loading was estimated as the equivalent of 50 birds being present all year with average inputs of 0.2 mg P/bird/yer and 0.95 mg N/bird/yr, typical literature rates; there were few data on waterfowl inputs, but a large decrease in bird use of the lake was noted after 2010. Direct septic system inputs were assumed to be negligible in light of sewering of the area around the lake.
The greatest difficulty came in matching in-lake phosphorus concentration and related variables (chlorophyll, Secchi transparency) to actual data, as the resulting sum of all loads translated into values higher than observed over the 2001-2011 period (Table 7). Given the abundance of rapidly settling particulate phosphorus and the uncertainty of how much of the internal load escapes the hypolimnion, the loads from the surface water sources (3 drainage areas) and the internal load were cut in half. This resulted in a load profile (Table 8) that produced a close match for actual phosphorus, chlorophyll and Secchi transparency data from the decade leading up to the USEPA study (Table 9). Nitrogen data are limited and incomplete, so little effort was devoted to calibrating nitrogen predictions, but they seem reasonable based on existing data.
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Table 5. Land Use Data for the Watershed of Lake Attitash
BASIN 1 BASIN 2 BASIN 3 TOTAL Attitash Back River SWW 2015 1998 LAND USE AREA (HA) AREA (HA) AREA (HA) AREA (HA) % % Urban 1 (LDR) 11.5 24.9 14.1 50.5 5.3 15 Urban 2 (MDR/Hwy) 55.2 15.8 2.0 73.1 7.7 5 Urban 3 (HDR/Com) 18.3 4.7 3.2 26.2 2.8 4 Urban 5 (P/I/R/C) 5.1 0.0 0.0 5.1 0.5 1 Open 3 (Urban) 0.0 37.7 0.0 37.7 4.0 Urban Sub-Total 90.1 83.2 19.3 192.7 20.2 25.0
Agric 2 (Row Crop) 0.2 56.5 11.2 68.0 7.1 16 Agric 3 (Grazing) 0.8 44.1 0.7 45.6 4.8 Open 2 (Meadow) 7.5 8.3 5.5 21.3 2.2 Agric. Sub-Total 8.6 108.9 17.3 134.8 14.1 16.0
Forest 1 (Upland) 88.8 359.5 61.0 509.3 53.5 55 Forest 2 (Wetland) 10.2 39.9 12.3 62.5 6.6 Open 1 (Wetland/Lake) 11.1 36.7 5.8 53.6 5.6 4 Natural Sub-Total 110.1 436.1 79.1 625.4 65.6 59.0
TOTAL 208.8 628.2 115.8 952.8 100.0 100.0
Table 6. Loading Features for the Three Drainage Areas of Lake Attitash
BASIN 1 BASIN 2 BASIN 3 Loading Feature Attitash Direct Back River SWW Water Load (cu.m/yr) 1081435 3017133 534533 Reality Check from Areal Yield X Basin Area 1009133 3036090 559493 Calculated Flow/Flow from Areal Yield 1.07 0.99 0.96 P Load (kg/yr) 74.5 177.5 25.8 Calculated P Concentration (mg/L) 0.069 0.059 0.048 Reality Check from Actual Data 0.068 0.060 0.047 Calculated Concentration/Measured Conc. 1.013 0.980 1.026 Calculated Basin Coefficient 0.36 0.28 0.22
Table 7. Total Phosphorus (ug/L) Based on 2001-2011 Data for the Lake Attitash System
Lake Lake Back Date Surface Bottom SW Inlet River Outlet Average 24 44 47 60 27 Median 19 38 35 49 19 Maximum 78 97 119 140 119 Minimum 5 18 5 6 10
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Table 8. Loading Summary for Current Conditions in Lake Attitash
LOAD SUMMARY Water Phosphorus Nitrogen DIRECT LOADS TO LAKE (CU.M/YR) (KG/YR) (KG/YR) ATMOSPHERIC 1698372 29.0 943.5 INTERNAL 0 117.0 234.0 WATERFOWL 0 10.0 47.5 SEPTIC SYSTEM 0 0.0 0.0 WATERSHED LOAD 4633102 138.9 5087.9
TOTAL LOAD TO LAKE 6331474 294.9 6313.0 (Watershed + Direct Loads) TOTAL INPUT CONC. (MG/L) 0.047 0.997
Table 9. Loading Comparison for Lake Attitash Management Scenarios
Internal Load Internal Maximum Removed SUMMARY TABLE FOR Background Load Feasible + Max SCENARIO TESTING Existing Conditions Conditions Removed BMPs BMPs Calibrated Model Actual Model Model Model Value Data Model Value Value Value Value Phosphorus (ppb) 23 24 9 15 19 10 Nitrogen (ppb) 600 700 400 581 448 430 Mean Chlorophyll (ug/L) 8.8 9.2 2.4 4.9 6.8 3.0 Peak Chlorophyll (ug/L) 30.1 30.6 9.2 17.4 23.6 10.9 Mean Secchi (m) 2.1 2.0 4.4 3.0 2.4 3.9 Peak Secchi (m) 4.1 4.0 5.3 4.6 4.3 5.1
Bloom Probability Probability of Chl >10 ug/L 30.7% 26.8 0.1% 4.8% 15.4% 0.4% Probability of Chl >15 ug/L 9.4% 19.5 0.0% 0.7% 3.4% 0.0% Probability of Chl >20 ug/L 2.9% 9.8 0.0% 0.1% 0.8% 0.0% Probability of Chl >30 ug/L 0.3% 2.4 0.0% 0.0% 0.1% 0.0% Probability of Chl >40 ug/L 0.1% 0 0.0% 0.0% 0.0% 0.0%
Once the model was calibrated to provide predictions that reasonably matched actual data values, it was used to evaluate various conditions and management scenarios (Table 9). To estimate the best possible condition of the pond, all developed and agricultural land uses were reset to forest and the internal load was reduced by 90%; this simulates pre-settlement conditions, except that a dam was added that increased lake area and volume, and those have not been changed. This “background” condition is not sustainably achievable by management, but provides a sense for the best possible condition the lake could have ever experienced.
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Sometimes, especially with large watersheds and shallow lakes, the background condition is not pristine at all, while for smaller watersheds and deeper lakes, the condition is often highly desirable. With a watershed to lake area ratio of just under 7:1 and a depth >9 m (30 feet), Lake Attitash would be expected to exhibit desirable conditions and does under this model scenario. The average phosphorus concentration is estimated to be 9 ug/L, mean Secchi transparency is predicted at 4.4 m (14.5 feet), and chlorophyll would very rarely exceed 10 ug/L. The lake has not experienced conditions like that in well over 100 years.
Working from the current conditions and reducing the internal load by 90%, something that can be achieved by multiple means, albeit at substantial cost, the lake is predicted to have an average phosphorus concentration of 15 ug/L, Secchi transparency of 3.0 m (10 feet), and chlorophyll in excess of 10 ug/L <5% of the time. These conditions represent a major improvement over the current condition and involve just control of internal loading of phosphorus.
Altering the current conditions to reflect maximum application of BMPs in the watershed but no in-lake work to control internal load, phosphorus would be expected to average 19 ug/L, Secchi transparency would average 2.4 m (7.9 feet), and chlorophyll would exceed 10 ug/L slightly more than 15% of the time. This represents some improvement over current conditions, but is based on getting the greatest reduction in watershed loading that seems feasible. Getting such a reduction would be a slow process and very expensive, based on experience elsewhere and literature examples. This scenario extends the type of work that has been done over the last two decades in the watershed, and would represent application of similar techniques to more locations and increasing cost. Watershed management should be part of any long-term plan to protect a lake, but diminishing returns on investment, ongoing maintenance needs, and practical limits to application generally prevent reductions of more than 50%. As substantial effort has already been put into watershed management for Lake Attitash, the expected reduction in this case is <50%, and is reflected in the predicted outcome within the lake.
Both the CDM (1999) and USEPA (2014) reports discuss watershed management needs on a fairly specific basis. Many of the CDM suggestions were followed up on and grants were received to help pay the capital cost of improvements. Very specific needs were itemized by the USEPA and include upgrade and maintenance actions relating to existing BMPs as well as new applications. Further watershed management is undoubtedly worthwhile, but appears unlikely to achieve universally acceptable conditions in Lake Attitash. In-lake action directed at the internal load will be needed as well, and by itself provides more improvement than the maximum feasible watershed management approach.
Combining the maximum feasible watershed approach and internal load reduction, the model predicts an average phosphorus level of 10 ug/L, Secchi transparency of 3.9 m (13 feet), and chlorophyll that only rarely exceeds 10 ug/L. This comes fairly close to achieving background conditions, but represents a very expensive and ongoing effort that may not be sustainable. Watershed management and internal load control are basically additive in this case; watershed management would limit the build-up of phosphorus reserves that fuel the internal load, but there is already so much phosphorus present that the linkage is not tight at all.
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In most cases the most thorough management package conceivable does not come so close to the background condition, but the small watershed size and very large potential internal load create more opportunity for improvement. Just how much improvement is needed is a subject for local discussion. Bear in mind that as the water becomes clearer, light penetration to a hospitable bottom will foster greater growth of rooted plants, a problem in this lake already in some areas. Some balance of the management of algae and plants may be warranted, and there will be multiple opinions on what constitutes acceptable water clarity. It seems safe to say that typical summer conditions are currently unacceptable, so some degree of improvement is needed.
Target Phosphorus Concentration The choice of a target phosphorus level is partly subjective, as described above; greater water clarity is obviously desirable, but increased rooted plant management needs are to be expected as clarity increases. The primary goal with regard to algae control would seem to be the elimination of cyanobacteria blooms, as these include floating scums and can produce taste and odor and toxicity. Moderate water clarity with other algae dominating would enhance the food web and improve fishing. The USEPA report describes the current fishery status as poor, with low abundance of larger gamefish and many small baitfish. The fishery can be managed by stocking, but improvement of water quality and management of rooted plants could trigger natural improvement while solving other lake problems.
Predicted background conditions are unlikely to be achieved in Lake Attitash, although the maximum feasible watershed effort and internal load control are predicted to come close to that “ideal” condition. Evaluation of historic phosphorus levels suggest that blooms are common at concentrations >20 ug/L, a common threshold in many southern New England lakes. Yet we are not really dealing with a threshold so much as a distribution, with conditions improving as the phosphorus concentration declines. Just how much to shift that distribution is the real question, with mean phosphorus level just representing the distribution.
Phosphorus criteria established by the USEPA for New England suggest values near 10 ug/L to avoid algae blooms and use impairment, but this has proven inadequate in cases where the algae grow at the sediment-water interface and then float upward to cause blooms in lakes with relatively low water column phosphorus levels. And in many cases it is cyanobacteria that use this bloom formation mode, suggesting that both water column phosphorus concentration and surficial sediment phosphorus availability must be lowered. This strongly favors including a technique to address internal loading in the management plan.
Some discussion among interested town parties and members of the Lake Attitash Association is warranted, but we can approximate a threshold by considering desired clarity and acceptable frequency of algae nuisances. As algae abundance is a distribution as described above, having no nuisances is not really a legitimate goal in almost any lake, but minimizing the frequency of algae above some target level is appropriate. A chlorophyll level of 10 ug/L is often applied, as values below 10 ug/L are associated with adequate clarity for most contact recreation but still represent adequate fertility to support a desirable fishery. However, as the ratio of chlorophyll to biomass is variable among algal groups, 10 ug/L of cyanobacteria chlorophyll may be a lot more
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biomass than for 10 ug/L of green algae chlorophyll. No threshold is perfect, but if chlorophyll remains under 10 ug/L 90% of the time, serious blooms will be unlikely.
Clarity relates directly to algae and the phytopigment chlorophyll for Lake Attitash, albeit with variability. Incoming storm water carries substantial non-algal particles, but settling appears rapid and most turbidity in the lake appears to be from algae. Secchi transparency is usually used as an indication of lake suitability for contact recreation. Values under about 1.2 m (4 feet) are usually associated with algae blooms and undesirable visibility, and used to be grounds for closing beaches in Massachusetts. Values >2 m (6.7 ft) are considered adequate for swimming, but most people prefer clarity near 3 m (10 feet). Values approaching 5 m (16.5 feet) are rare in Massachusetts lakes in summer, and would be considered outstanding clarity. The current average for Lake Attitash is close to 2 m, but values in summer are often closer to 1 m. A value close to 3 m would seem like an appropriate target for Secchi transparency.
Translating a 90th percentile chlorophyll value of 10 ug/L and an average Secchi transparency of 3 m into a target phosphorus concentration with the model suggests a value between 15 and 17 ug/L. Any phosphorus concentration within or below that range should provide desirable conditions based on the proposed criteria. If different levels of clarity and chlorophyll are preferred, a different phosphorus threshold can be calculated, but this seems to be a reasonable starting point for discussion.
Referring to Table 9, a 90% reduction of internal loading meets the goal, while maximum practical application of BMPs does not. The combination of more BMPs in the watershed plus internal load reduction is more than adequate, so control of internal loading with continued attention to watershed loading as opportunities present themselves seems advisable.
Management Options Review There is a wide variety of options for managing algae, ranging from watershed techniques to limit the input of nutrients to physical techniques such as flushing or dredging to chemical techniques such as algaecides or phosphorus inactivation to biological techniques such as food web manipulations to maximize grazing (Table 10). These techniques attempt to either reduce growth or increase loss of algae, and the duration of benefits varies widely by technique. If external loading is low, dredging or phosphorus inactivation may provide decades of improved conditions. If external loading remains high, repetitive algaecide applications or near constant input of air or oxygen may be necessary to minimize blooms. The key is to match techniques to the situation with consideration of technical feasibility and expected results, economic affordability and longer term sustainability, and institutional acceptability, which includes everything from acceptance by the lake association to permits from the town and state agencies.
High algal productivity is not necessarily undesirable; if algae are efficiently processed in the food web, biomass will not build up and more fish will be produced while acceptable water clarity is maintained. However, it is rare to have a biological structure that can process as much algae as can be produced at high nutrient levels, so keeping nutrient levels in a low to moderate range is usually a primary goal. Lowering the overall fertility of the system is preferred where feasible.
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Table 10. Options for Control of Algae and Floating Plants (Adapted from Wagner 2001)
OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES WATERSHED CONTROLS 1) Management for Includes wide range Acts against the May involve nutrient input of watershed and original source of considerable lag reduction lake edge activities algal nutrition time before intended to eliminate Creates sustainable improvement nutrient sources or limitation on algal observed reduce delivery to growth May not be lake May control sufficient to Essential component delivery of other achieve goals of algal control unwanted pollutants without some strategy where to lake form of in-lake internal recycling is Facilitates management not the dominant ecosystem Reduction of nutrient source, and management overall system desired even where approach which fertility may internal recycling is considers more than impact fisheries important just algal control May cause shift in nutrient ratios which favor less desirable algae
1a) Point source More stringent Often provides May be very controls discharge major input expensive in requirements reduction terms of capital May involve Highly efficient and operational diversion approach in most costs May involve cases May transfer technological or Success easily problems to operational monitored another watershed adjustments Variability in May involve results may be pollution prevention high in some plans cases 1b) Non-point Reduction of sources Removes source May require source of nutrients Limited ongoing purchase of land controls May involve costs or activity elimination of land May be viewed as uses or activities that limitation of release nutrients “quality of life” May involve Usually requires alternative product education and use, as with no gradual phosphate fertilizer implementation
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OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES 1c) Non-point source Capture of pollutants Minimizes Does not address pollutant between source and interference with actual sources trapping lake land uses and May be expensive May involve activities on necessary drainage system Allows diffuse and scale alteration phased May require Often involves implementation substantial wetland treatments throughout maintenance (det./infiltration) watershed May involve storm Highly flexible water collection and approach treatment as with Tends to address point sources wide range of pollutant loads IN-LAKE PHYSICAL CONTROLS 2) Circulation and Use of water or air Reduces surface May spread destratification to keep water in build-up of algal localized impacts motion scums May lower Intended to prevent May disrupt growth oxygen levels in or break of blue-green algae shallow water stratification Counteraction of May promote Generally driven by anoxia improves downstream mechanical or habitat for impacts pneumatic force fish/invertebrates Can eliminate localized problems without obvious impact on whole lake 3) Dilution and Addition of water of Dilution reduces Diverts water flushing better quality can nutrient from other uses dilute nutrients concentrations Flushing may Addition of water of without altering wash desirable similar or poorer load zooplankton from quality flushes Flushing minimizes lake system to minimize detention; response Use of poorer algal build-up to pollutants may quality water May have be reduced increases loads continuous or Possible periodic additions downstream impacts
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OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES 4) Drawdown Lowering of water May reduce Possible impacts over autumn period available nutrients on non-target allows oxidation, or nutrient ratios, resources desiccation and affecting algal Possible compaction of biomass and impairment of sediments composition water supply Duration of Opportunity for Alteration of exposure and degree shoreline clean- downstream of dewatering of up/structure repair flows and winter exposed areas are Flood control utility water level important May provide rooted May result in Algae are affected plant control as greater nutrient mainly by reduction well availability if in available flushing nutrients. inadequate 5) Dredging Sediment is Can control algae if Temporarily physically removed internal recycling is removes benthic by wet or dry main nutrient invertebrates excavation, with source May create deposition in a Increases water turbidity containment area for depth May eliminate dewatering Can reduce fish community Dredging can be pollutant reserves (complete dry applied on a limited Can reduce dredging only) basis, but is most sediment oxygen Possible impacts often a major demand from containment restructuring of a Can improve area discharge severely impacted spawning habitat Possible impacts system for many fish from dredged Nutrient reserves are species material disposal removed and algal Allows complete Interference with growth can be renovation of recreation or limited by nutrient aquatic ecosystem other uses during availability dredging 5a) “Dry” excavation Lake drained or Tends to facilitate a Eliminates most lowered to very thorough aquatic biota maximum extent effort unless a portion practical May allow drying left undrained Target material dried of sediments prior Eliminates lake to maximum extent to removal use during possible Allows use of less dredging Conventional specialized excavation equipment equipment used to remove sediments
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OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES 5b) “Wet” excavation Lake level may be Requires least Usually creates lowered, but preparation time or extreme turbidity sediments not effort, tends to be Normally requires substantially least cost dredging intermediate exposed approach containment area Draglines, bucket May allow use of to dry sediments dredges, or long- easily acquired prior to hauling reach backhoes used equipment May disrupt to remove sediment May preserve ecological aquatic biota function Use disruption 5c) Hydraulic removal Lake level not Creates minimal Often leaves reduced turbidity and some sediment Suction or impact on biota behind cutterhead dredges Can allow some Cannot handle create slurry which lake uses during coarse or debris- is hydraulically dredging laden materials pumped to Allows removal Requires containment area with limited access sophisticated and Slurry is dewatered; or shoreline more expensive sediment retained, disturbance containment area water discharged 6) Light-limiting dyes Creates light Creates light limit May cause and surface covers limitation on algal growth thermal without high stratification in turbidity or great shallow ponds depth May facilitate May achieve some anoxia at control of rooted sediment plants as well interface with water 6.a) Dyes Water-soluble dye is Produces appealing May not control mixed with lake color surface bloom- water, thereby Creates illusion of forming species limiting light greater depth May not control penetration and growth of shallow inhibiting algal water algal mats growth Altered thermal Dyes remain in regime solution until washed out of system.
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OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES 6.b) Surface covers Opaque sheet Minimizes Minimizes material applied to atmospheric and atmospheric gas water surface wildlife pollutant exchange inputs Limits recreation 7) Mechanical removal Filtering of pumped Algae and Filtration requires water for water associated nutrients high backwash supply purposes can be removed and sludge Collection of from system handling floating scums or Surface collection capability mats with booms, can be applied as Labor and/or nets, or other needed capital intensive devices May remove Variable Continuous or floating debris collection multiple applications Collected algae dry efficiency per year usually to minimal volume Possible impacts needed on non-target aquatic life 8) Selective Discharge of bottom Removes targeted Possible withdrawal water which may water from lake downstream contain (or be efficiently impacts of poor susceptible to) low May prevent anoxia water quality oxygen and higher and phosphorus May promote nutrient levels build up in bottom mixing of May be pumped or water remaining poor utilize passive head May remove initial quality bottom differential phase of algal water with blooms which start surface waters in deep water May cause May create unintended coldwater drawdown if conditions inflows do not downstream match withdrawal 9) Sonication Sound waves disrupt Supposedly affects Unknown effects algal cells only algae (new on non-target technique) organisms Applicable in May release localized areas cellular toxins or other undesirable contents into water column
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OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES 10) Hypolimnetic Addition of air or Oxic conditions May disrupt aeration or oxygen provides reduce P thermal layers oxygenation oxic conditions availability important to fish Maintains Oxygen improves community stratification habitat Theoretically Can also withdraw Oxygen reduces promotes water, oxygenate, build-up of reduced supersaturation then replace cpds with gases harmful to fish IN-LAKE CHEMICAL CONTROLS 11) Algaecides Liquid or pelletized Rapid elimination Possible toxicity algaecides applied to of algae from water to non-target target area column , normally species Algae killed by with increased Restrictions on direct toxicity or water clarity water use for metabolic May result in net varying time after interference movement of treatment Typically requires nutrients to bottom Increased oxygen application at least of lake demand and once/yr, often more possible toxicity frequently Possible recycling of nutrients 11a) Forms of copper Cellular toxicant, Effective and rapid Possible toxicity disruption of control of many to aquatic fauna membrane transport algae species Accumulation of Applied as wide Approved for use in copper in system variety of liquid or most water supplies Resistance by granular certain green and formulations blue-green nuisance species Lysing of cells releases nutrients and toxins
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OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES 11b) Peroxides Disrupts most Rapid action Much more cellular functions, Oxidizes cell expensive than tends to attack contents, may limit copper membranes oxygen demand and Limited track Applied as a liquid toxicity record or solid. Possible recycling Typically requires of nutrients application at least once/yr, often more frequently
11c) Synthetic organic Absorbed or Used where copper Non-selective in algaecides membrane-active is ineffective treated area chemicals which Limited toxicity to Toxic to aquatic disrupt metabolism fish at fauna (varying Causes structural recommended degrees by deterioration dosages formulation) Rapid action Time delays on water use 12) Phosphorus Typically salts of Can provide rapid, Possible toxicity inactivation aluminum, iron or major decrease in to fish and calcium are added to phosphorus invertebrates, the lake, as liquid or concentration in especially by powder water column aluminum at low Phosphorus in the Can minimize pH treated water column release of Possible release is complexed and phosphorus from of phosphorus settled to the bottom sediment under anoxia or of the lake May remove other extreme pH Phosphorus in upper nutrients and May cause sediment layer is contaminants as fluctuations in complexed, reducing well as phosphorus water chemistry, release from Flexible with especially pH, sediment regard to depth of during treatment Permanence of application and Possible binding varies by speed of resuspension of binder in relation to improvement floc in shallow redox potential and areas pH Adds to bottom sediment, but typically an insignificant amount
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OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES 13) Sediment Addition of Can reduce Possible impacts oxidation oxidants, binders phosphorus supply on benthic biota and pH adjustors to to algae Longevity of oxidize sediment Can alter N:P ratios effects not well Binding of in water column known phosphorus is May decrease Possible source of enhanced sediment oxygen nitrogen for blue- Denitrification is demand green algae stimulated 14) Settling agents Closely aligned with Removes algae and Possible impacts phosphorus increases water on aquatic fauna inactivation, but can clarity without Possible be used to reduce lysing most cells fluctuations in algae directly too Reduces nutrient water chemistry Lime, alum or recycling if floc during treatment polymers applied, sufficient Resuspension of usually as a liquid or Removes non-algal floc possible in slurry particles as well as shallow, well- Creates a floc with algae mixed waters algae and other May reduce Promotes suspended particles dissolved increased Floc settles to phosphorus levels sediment bottom of lake at the same time accumulation Re-application typically necessary at least once/yr 15) Selective nutrient Ratio of nutrients Can reduce algal May result in addition changed by levels where greater algal additions of selected control of limiting abundance nutrients nutrient not feasible through uncertain Addition of non- Can promote non- biological limiting nutrients nuisance forms of response can change algae May require composition of algal Can improve frequent community productivity of application to Processes such as system without maintain desired settling and grazing increased standing ratios can then reduce algal crop of algae Possible biomass downstream effects
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OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES IN-LAKE BIOLOGICAL CONTROLS 16) Enhanced grazing Manipulation of May increase water May involve biological clarity by changes introduction of components of in algal biomass or exotic species system to achieve cell size without Effects may not grazing control over reduction of be controllable or algae nutrient levels lasting Typically involves Can convert May foster shifts alteration of fish unwanted algae into in algal community to fish composition to promote growth of Harnesses natural even less grazing zooplankton processes desirable forms 16.a) Herbivorous fish Stocking of fish that Converts algae Typically requires eat algae directly into introduction of potentially non-native harvestable fish species Grazing pressure Difficult to can be adjusted control over long through stocking term rate Smaller algal forms may be benefited and bloom 16.b) Herbivorous Reduction in Converts algae Highly variable zooplankton planktivorous fish to indirectly into response promote grazing harvestable fish expected; pressure by Zooplankton temporal and zooplankton response to spatial variability May involve increasing algae may be high stocking piscivores can be rapid Requires careful or removing May be monitoring and planktivores accomplished management May also involve without action on 1-5 yr stocking introduction of non- basis zooplankton or native species Larger or toxic establishing refugia Generally algal forms may compatible with be benefitted and most fishery bloom management goals
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OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES 17) Bottom-feeding Removes fish that Reduces turbidity Targeted fish fish removal browse among and nutrient species are bottom deposits, additions from this difficult to control releasing nutrients to source Reduction in fish the water column by May restructure populations physical agitation fish community in valued by some and excretion more desirable lake users manner (human/non- human) 18) Microbial Addition of Shifts nutrient use Minimal competition microbes, often with to organisms that scientific oxygenation, can tie do not form scums evaluation up nutrients and or impair uses to N control may limit algal growth same extent as still favor Tends to control N algae cyanobacteria more than P Harnesses natural May need processes aeration system to May decrease get acceptable sediment results 19) Pathogens Addition of May create Largely inoculum to initiate lakewide experimental attack on algal cells “epidemic” and approach at this May involve fungi, reduction of algal time bacteria or viruses biomass May promote May provide resistant nuisance sustained control forms through cycles May cause high Can be highly oxygen demand specific to algal or release of group or genera toxins by lysed algal cells Effects on non- target organisms uncertain 20) Competition and Plants may tie up Harnesses power of Some algal forms allelopathy by plants sufficient nutrients natural biological appear resistant to limit algal growth interactions Use of plants may Plants may create a May provide lead to problems light limitation on responsive and with vascular algal growth prolonged control plants Chemical inhibition Use of plant of algae may occur material may through substances cause depression released by other of oxygen levels organisms
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OPTION MODE OF ACTION ADVANTAGES DISADVANTAGES 20a) Plantings for Plant growths of Productivity and Vascular plants nutrient control sufficient density associated habitat may achieve may limit algal value can remain nuisance densities access to nutrients high without algal Vascular plant Plants can exude blooms senescence may allelopathic Can be managed to release nutrients substances which limit interference and cause algal inhibit algal growth with recreation and blooms Portable plant provide habitat The switch from “pods” , floating Wetland cells in or algae to vascular islands, or other adjacent to the lake plant domination structures can be can minimize of a lake may installed nutrient inputs cause unexpected or undesirable changes 20b) Plantings for Plant species with Vascular plants can Floating plants light control floating leaves can be more easily can be a shade out many algal harvested than most recreational growths at elevated algae nuisance densities Many floating Low surface species provide mixing and waterfowl food atmospheric contact promote anoxia 20c) Addition of Input of barley straw Materials and Success appears barley straw can set off a series of application are linked to chemical reactions relatively uncertain and which limit algal inexpensive potentially growth Decline in algal uncontrollable Release of abundance is more water chemistry allelopathic gradual than with factors chemicals can kill algaecides, limiting Depression of algae oxygen demand and oxygen levels Release of humic the release of cell may result substances can bind contents Water chemistry phosphorus may be altered in other ways unsuitable for non-target organisms
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Running through the list of options in Table 10, some techniques are less applicable than others. Dilution or flushing would require a tremendous amount of water with no obvious source; detention time must be kept under 3 weeks for flushing to work, and a supply of very clean water for dilution is absent. Additionally, internal load would overcome any dilution effect.
Drawdown was not recommended by the USEPA on the basis of hydrologic impacts, and it should be noted that the plants causing the greatest nuisances in recent years appear to be seed producing annuals resistant to drawdown control. From the perspective of algae control, drawdown might be useful for removing the worst quality water in fall after a summer of internal loading, but it will take decades to get any measureable improvement by that approach.
Surface covers would greatly restrict use of the lake, and dyes are permitted like algaecides; concentrations high enough to make a difference would not likely be allowed to flow downstream, and surface cyanobacteria scums may not be prevented.
Mechanical removal is not practical for microscopic algae. There is a portable diffused air flotation unit in development, and it might allow removal of microscopic algae, but no practical demonstration has yet occurred.
Sonication works on some algae but not others, and a large number of units would be needed to maintain control. Although solar options exist, these units are normally run by direct electricity and running power for all those units in the lake would be a major issue.
Sediment oxygenation could be beneficial, but this has not been a practiced technique for many years following field trials in the 1980s. Likewise, addition of settling agents for algae could work, but is not nearly as beneficial as aluminum, which inactivates phosphorus as well as settling algae. Additionally, adding a large mass of algae to the bottom would increase oxygen demand and may accelerate internal loading.
Selective nutrient addition is a fascinating area of study. It is possible to prolong a spring diatom dominated assemblage into summer by adding silica, and addition of nitrate could prevent cyanobacteria from becoming dominant. While having other algae dominant would represent an improvement and may help the food web in the lake, there would still be algae blooms and clarity would be low. Better options exist.
Removal of bottom feeding fish may provide some benefits, but would not impact internal or external loading to a great degree in this case. Microbial competition is a controversial discipline with literally no peer reviewed papers upon which to base rational conclusions. There are success stories, but there are also major failures. Enhanced microbial action may digest more organic sediment, and this generally requires oxygenation and some enzymes to break down certain hydrocarbons, but again there is no clear track record of success. Addition of pathogens has enjoyed two bursts of research over the last 4 decades, but no commercial products are currently available.
The use of plant or plant matter to discourage algae blooms is another interesting area of study. Decaying barley straw produces natural algaecides that seem to be most effective on
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cyanobacteria, but it is not a registered algaecidal product, so no commercial applicator can use it for that purpose. Abundant vascular plants, especially those with floating leaves, can restrict algae growth through shading, nutrient competition, or release of allelopathic substances, but Lake Attitash is too deep to grow a large enough plant community and if it did, the problems with plant biomass would be every bit as troublesome as algae blooms.
This narrows the field of viable candidate management methods to: 1. Watershed management for reduced phosphorus input: This involves source controls and/or pollutant trapping that lowers the external load to the lake. The CDM (1999) report focused on watershed BMPs and suggested a filter curtain as used at one Wachusett Reservoir inlet to enhance detention at the Back River inlet. A Gunderboom was added in the mid 2000's with some success in settling and retention behind the curtain material. It was removed and cleaned once in about 2011 and is still functioning. The USEPA (2014) report details specific watershed management needs and locations, including restoration of wetland function, maintenance of existing detention structures, and construction of storm water drainage system improvements that would enhance discharged water quality.
Most of the suggested activities have been pursued, with grants obtained to support these efforts. A Section 319 grant was obtained in 2001 and used to install BMPs in the Lake Shore Drive area (Amesbury, MA 2006). Those BMPs consisted of baffled tanks and deep sump catch basins to capture sediment and related contaminants that were freely entering the lake. Monitoring under a QAPP demonstrated the resulting improvement. A public outreach and education program helped inform residents of their role in minimizing pollution of the lake. This project became a template for Amesbury in addressing other watershed problems.
A second Section 319 grant was obtained in 2011 and used to pursue storm water management in additional areas. Issues on West Shore Road, Ahern Circle, Lakeshore Drive, Merrill Avenue, Meadow Avenue, Attitash Avenue, First Street, Lake Avenue and Birchmeadow Road were addressed using deep sump catch basins, leaching catch basins, detention ponds, rain gardens, vegetated swales, and buffer strips (Amesbury, MA 2014). The improvement in stormwater quality has been documented, but overall improvement of Lake Attitash has lagged, undoubtedly as a function of ongoing internal loading related to historical inputs to the lake. Action to protect the lake from further damage from the watershed has been substantial, but in-lake action is now needed to actually restore the lake.
The LLRM model for Lake Attitash suggests that even the most aggressive but feasible effort toward watershed management will not result in the proposed level of improvement in the lake itself, and indeed that has been the case. Watershed management should indeed be pursued further, but it will not be sufficient by itself. At the cost of needed improvements, which increases after the simpler actions are taken, local leadership will need to be selective and opportunistic to make any measureable difference in the incoming water quality.
2. Dredging: Removal of soft sediment would reduce the internal load and remove resting stages of both algae and rooted plants. On a large scale, dredging could set the lake back in time and greatly enhance conditions. The areas that need to be dredged for algae control do not completely overlap with areas in need of rooted plant control, and the cost would be very
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high to dredge adequately to make a difference. Focusing on algae control, the primary purpose of this assessment, and considering the results of the sediment testing (Table 4), at least 10 cm (4 inches) of sediment would have to be removed (and probably much more) to substantially reduce internal loading.
Over the 194 acre area that experiences summer anoxia, that equates to 64 acre feet of sediment, or slightly more than 100,000 cubic yards. At a low end cost of $30/cy, the cost would be at least $3 million, and it would not be surprising for the cost to be 3-5 times as much based on additional quantity or additional disposal cost. Just the necessary testing, engineering, and permitting for a dredging project can expected to cost in excess of $100,000 in a case like this. As depth recovery is not really an issue in deeper water, the cost does not justify the expense when other options exist.
3. Enhanced grazing on algae: At the levels of phosphorus observed in Lake Attitash, restructuring the biological community to maximize consumption of algae does have merit. The key is to promote the largest possible population of a large bodied Daphnia, a grazing zooplankter that can clear the water of most algae. Some of the cyanobacteria will be difficult to graze due to particle size and possible toxicity, but encouraging Daphnia is appropriate. However, the zooplankton community has experienced a decline in Daphnia in recent years (Ruggirio 2014) and testing in August of 2015 found no Daphnia. Smaller cladocerans such as Diaphanosoma contributed to a moderate biomass of 59 ug/L, but provide inadequate grazing capacity.
The USEPA (2014) description of the fish community is consistent with these findings. Enhancement would involved altering the fish community to favor Daphnia and patiently awaiting a response. Stocking gamefish or conducting a laborious panfish removal project would be necessary, both as substantial cost and with no guarantee of results. Biological approaches tend to carry substantial variability and require vigilance and maintenance to provide continued benefits. While improvement of the fish community with a possible beneficial cascading effect on zooplankton is highly desirable, it might be better to control algae more directly and monitor for natural improvement first. Additionally, a MET Grant was applied for in 2011 to add larger game fish but did not get funded despite EPA, MA Fish & Game and UNH support.
4. Algaecides: Application of algaecides can be effective and despite potential noon-target impacts, they remain a valuable tool for algae control. Even the best management for nutrient control may be inadequate on occasion, as with catastrophic events (floods or fire) or the impact of climate change (very warm summer periods). In such cases a well-timed algaecide treatment can be effective at reasonable cost. The main issue is timing; monitoring must be adequate to detect the early stages of a bloom, so that algaecide treatment prevents a bloom, not destroys it once formed.
Killing off a major bloom carries risks of higher oxygen demand, toxin release at dangerous levels, and other water quality impacts that have resulted in regulation of cyanobacteria treatment in Massachusetts. Copper and peroxide are the main algaecides in use, with copper
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less expensive. Having this option available is worthwhile, but it does not need to be the first line of defense in Lake Attitash.
5. Circulation: Keeping the water in motion from top to bottom in a lake, helps satisfy the oxygen demand, can limit nutrient release from sediment, and may disrupt the growth of some algae, notably the buoyant cyanobacteria that cause many of the blooms in Lake Attitash. There are several limitations to this approach, however. It is difficult to mix to the sediment-water interface without stirring up sediment, so a small anoxic zone may persist and allow release of phosphorus that may then be moved upward by the circulation system and aid algae growth. It is hard to overcome the input of heat from the sun during a prolonged summer dry spell, so circulation systems must either be greatly overpowered to handle all situations or will fail during those dry spells. Whatever nutrients are in the lake will be continually mixed and made available to algae, and with continued watershed loading, this may be enough to promote blooms even if the internal load is reduced.
If the water can be mixed to a depth of at least 9 m (>30 feet), the time spent in darkness may reduce growth, but much of the lake is not >9 m deep. Loss of the hypolimnion through mixing may have ecological consequences, although the current lack of oxygen during summer minimizes the habitat value of this zone. The capital cost in not minor (expect up to $2000 per acre addressed, with 194 acres to be considered), and there is an ongoing operational cost for power and maintenance that is not trivial (expect $100-$200 per acre per year). While potentially beneficial, this approach has too many drawbacks in this case to be recommended.
6. Oxygenation: Adding oxygen to deep water is not always necessary, but almost always provides water quality benefits. As long as anoxia can be prevented, release of iron-bound phosphorus can be minimized and a variety of other undesirable interactions between sediment and water can be limited.
Aside from circulation to distribute oxygenated surface waters to near the bottom, there are four ways to add oxygen directly to target waters without breaking stratification. The most efficient method, diffused oxygen, is best applied to a hypolimnion of at least 6 m (20 feet) in thickness to allow enough vertical distance for the oxygen to be absorbed and avoid destratifying the lake. At the deepest point with the strongest stratification in Lake Attitash, there are barely 6 m of hypolimnion thickness, and most of the target volume is thinner.
The use of hypolimnetic aeration chambers, which circulate air in a chamber to oxygenate water that is then released back into the hypolimnion, can be effective but is very inefficient; oxygen transfer from air is slow, so a lot of it never gets transferred and more water and air have to be moved to achieve the desired results at substantial power cost.
Speece cones use pure oxygen in a submerged chamber, increasing efficiency at a large capital cost, but would be applicable. The main issue would be creating a platform for the equipment on the very soft sediment in deep water; some dredging may be necessary, greatly increasing cost and permitting delay.
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Most intriguing is the sidestream supersaturation approach, in which water is pumped from the lake to a land-based pressurized chamber where pure oxygen is add to get a supersaturated solution that is put back into the hypolimnion. Less water has to be moved, although the cost is still substantial. Either the sidestream supersaturation approach or diffused oxygen, with the acknowledgement that some destratification may occur. is most appropriate for Lake Attitash.
7. Phosphorus inactivation: Inactivation of phosphorus has been practiced in New England for over 30 years, and this technique has been refined and advanced considerably over those decades. Presentations as recently as November 2015 at the NALMS conference broke new ground on understanding key processes and advancing treatment effectiveness, so those considering phosphorus inactivation need to be up to date on the associated science.
There are 3 ways to use phosphorus inactivation: Treatment of surficial sediment with larger doses to prevent release of phosphorus, mainly from iron subjected to anoxia Treatment of the water column with lower doses to inactivate and settle phosphorus and algae Treatment of inflows with lower doses, mainly during storms, to inactivate incoming phosphorus and settle associated solids Sediment treatments can involve calcium, aluminum or lanthanum, with aluminum having the longest and best documented track record. Sediment treatments in deeper lakes tend to provide benefits for about 20 years, as iron-bound phosphorus is effectively neutralized and that is the main source of phosphorus through internal loading. Sediment treatments in deeper lakes tend to provide benefits for about 20 years, with termination of benefits largely linked to upward migration of iron-bound phosphorus through the inactivation zone (usually the upper 4-10 cm of sediment).
One alternative that may be worth investigating is Phoslock, which is a mix of bentonite clay and lanthanum. It is not yet approved for use in Massachusetts, but should be before long. The lanthanum binds well with phosphorus and the clay helps seal the surficial sediment. While the track record is too short to make a firm prediction, the potential to both reduce phosphorus availability and lower oxygen demand by covering the organic sediment with a thin clay layer is appealing. It seems unlikely that the normal dose will be adequate to cover the soft sediment, and any dose may eventually sink into that very loose sediment, but the company that markets Phoslock could conduct tests at limited cost. However, the cost of Phoslock is several times that of aluminum application.
Application of lower doses of aluminum to the water column can reduce available phosphorus and at least temporarily lower algae biomass. Water column treatment has been used elsewhere with varied but not lasting results. This is entirely consistent with the assumed mode of internal loading, whereby a very large supply of sediment phosphorus can be released and replace inactivated water column phosphorus within the same summer season. Higher doses or more frequent treatments may provide better or longer lived results; certainly low dose treatments would fare better if the sediment was first treated with a larger
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dose, but that one large dose to inactivate phosphorus in surficial sediment may be all that is needed in Lake Attitash.
Inactivation of incoming storm water phosphorus involves a dosing station delivering an approximate dose whenever it rains. This can be done manually or automatically at slightly higher cost, and has been very successful in many cases in Florida and at the only active installation in Massashusetts (Morses Pond in Wellesley). A dose of 1-3 mg/L is targeted, with pumps and chemical storage sized accordingly. Systems are commercially available or can be custom made at a cost on the order of $100,000 per station. Treatment is generally conducted in the spring to ensure that conditions are optimal going into summer, but continued summer treatment can be conducted if warranted. The Back River might be a goodcandidate for such a dosing station, and would not need to be operated all the time (during most summer seasons the flow in to Back River becomes minimal), but again the single inactivating dose to the sediment may be all that is needed in this case. A dosing station on a tributary would be more of an alternative to further watershed management than a remedy for internal loading. Logically, treatment behind the gunderboom during storm events would be the most advantageous approach. Annual operational cost is largely a matter of necessary aluminum chemical supply.
Phosphorus inactivation provides great flexibility of application to Lake Attitash. Application of a dose sufficient to inactivate iron-bound phosphorus in the surficial sediment would greatly reduce internal loading, which based on model results, could achieve the desired conditions in the lake. Additional dosing at an inlet or in the lake could counter ongoing watershed loading as warranted.
Oxygenation Potential Oxygenation represents an option for curtailing internal loading while enhancing deep water habitat. There are four methods to oxygenate without disrupting stratification, and two of these seem most suited to Lake Attitash. Diffused oxygen injection may cause some destratification, but is a very simple approach with minimal moving parts or power requirements (Figure 14). Liquid oxygen is stored in a tank. Opening the valve allows the liquid to move into a vaporizer, and the gas then travels through hoses to the target area under its own gas pressure. Porous hose in the target volume allows small oxygen bubbles to escape. The bubbles are absorbed by the low oxygen water as the bubbles rise.
The recommended alternative oxygen method is sidestream supersaturation (Figure 15). Water is pumped into the oxygenation chamber on land and oxygen is added under pressure, allowing much more oxygen to be absorbed by the water. The superoxygenated water is returned to the target zone, where it spreads and oxygen additionally diffused into lower oxygen water. This technique is perhaps more suitable for thin hypolimnions than the diffused oxygen approach, but has so far been more expensive. However, this technique is still gaining in popularity and costs may decline in the near future.
A range of $2000-4000/ac for capital cost is appropriate (Wagner 2014), which for 194 acres amounts to a capital expense of $388,000 to $776,000. As anoxia may cover only 130 acres in
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Figure 14. Land Based (left) and Underwater (right) Elements of a Diffused Oxygen System (Courtesy of Mobley Engineering)
Figure 15. Schematic of a Sidestream Supersaturation Oxygenation System (Courtesy of Oklahoma Water Resources Board)
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the summer, a smaller system might be installed for $260,000 to $520,000. Annual operation can be estimated at $400-600/ac by one method (with oxygen or power as the main costs), or $77,600 to $116,400 for the 194 acre target area and $52,000 to $78,000 for the 130 acre area. However, the per acre cost estimation approach is less reliable for thinner hypolimnions, especially if oxygen demand is not excessive.
The alternative method for cost estimation for oxygenation is to consider the actual amount of oxygen needed and a cost per kg of oxygen delivered (Wagner 2014). The oxygen demand is about 1.1 g/m2/day, and can be expected to increase by 50 to 100% when oxygen is supplied, so about 1.65 to 2.2 g/m2/day will be needed. Over 194 acres, that equates to about 1300 to 1728 kg/day. Appropriate delivery systems cost $200 to $1000 per kg, so the capital cost could range from $260,000 to $1.7 million. Operating cost could range from $0.25 to $2/kg; assuming operation for 60-90 days, a cost between $20,000 and $311,000 can be estimated.
These estimates could be scaled down to 130 acres, the minimum summer anoxic area, at a 33% savings. However, the estimates represent very wide ranges for both capital and operational costs, and further refinement is needed by an appropriate design firm if this approach is to be pursued. The information for systems now in place suggests that application to Lake Attitash would be on the small end of the encountered range, and would not enjoy an economy of scale that accounts for the low end of the cost range. A capital cost of not less than $700,000 and an operational cost of not less than $100,000/year should be assumed.
While oxygenation offers potentially great benefits to Lake Attitash, the wide range of capital cost and the ongoing and substantial annual operational cost are negative factors. Additionally, there is a land-based footprint which will require space somewhere near the lake that can accommodate large tanker truck access. This may be logistically problematic in this case.
Phosphorus Inactivation Potential Phosphorus inactivation involves the binding of phosphorus by added compounds that make it unavailable for uptake by algae. Treatment of incoming water, lake water or lake sediment is possible and applicable to Lake Attitash, but the primary need appears to be inactivation of iron- bound phosphorus in surficial sediments that provides a substantial internal load (Figure 16). Treatment of incoming storm water or the lake water column could be considered, but each would involve repetitive treatments over years, while a single adequate dose to the sediment over an area of about 194 acres could minimize internal loading for up to 20 years and meet proposed lake target conditions.
Aluminum has been the phosphorus binder of choice in Massachusetts for the last 20 years. Aluminum sulfate can be applied by itself where alkalinity is high, but in most cases with a high dose to inactivate sediment phosphorus, sodium aluminate is applied with the aluminum sulfate to keep the pH stable. Polyaluminum chloride is gaining popularity for inflow and water column treatments. There is another binder that has been used in recent years called Phoslock, which is bentonite clay with lanthanum attached to binding sites. It may do a superior job capturing phosphorus from the water column, and may do as good a job on surficial sediments, but at a
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Figure 16. Application of Aluminum (barge applying aluminum on the left, vertical gradient of floc formation on the right) cost of about five times the cost of an appropriate aluminum dose. Given the cost factor and experience with aluminum, dosing the sediments with aluminum is recommended.
Successful aluminum treatment is a function of supplying an adequate dose to the appropriate treatment area. It is generally acknowledged that the targeted treatment area should be the area of sediment that can experience anoxia, which facilitates the release of phosphorus bound by iron (Fe-P). The necessary dose is a matter of both the Fe-P concentration and other sediment constituents that may compete with Fe-P for binding sites on the applied aluminum compounds. This is an area of current study that has some degree of uncertainty attached to it.
The aluminum to phosphorus ratio (Al:P) necessary for effective inactivation varies inversely with Fe-P concentration, as lower Fe-P levels means that other constituents are abundant and
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compete for binding sites (James and Bischoff 2015). When Fe-P is high it tends to occupy more of the binding sites and the necessary ratio of Al to P increases as Fe-P declines. The range of Al:P ratios for successful treatments tends to range from 10 to 150, although treatments at >100 g/m2 have not been needed in MA. The applied range in MA is 10 to 100 g/m2, but most treatments have been between 30 and 75 g/m2 and the average is close to 50 g/m2. Treated lakes with higher Fe-P had a lower Al:P ratio, while those with lower Fe-P required a higher Al:P ratio.
Fe-P levels in most of Lake Attitash would be considered moderate (Table 4). Based on equations from James and Bischoff (2015), the average Fe-P concentration suggests a 40:1 Al:P ratio, which would yield a dose of 209 g/m2 for the whole target area, much higher than any MA treatment to date. Dose determination by stoichiometry calculations suggests 117 g/m2 for the 20 ac with high Fe-P and 49 g/m2 for other areas. The Al:P ratio is 10:1 in these calculations, at the low end of the recommended range. If we simply apply the low end Al:P ratio of 10:1 and the average Fe-P concentration from tested samples, the dose should be 52 g/m2, similar to that of many other MA treatments.
Out of concern over the variability and uncertainty of calculated doses, we developed a laboratory aluminum assay about a decade ago. In those assays, small amounts of sediment are treated with aluminum in the lab then retested for Fe-P. As the dose rises, remaining Fe-P declines, but almost never in a linear pattern. This allows visualization of how well Fe-P can be inactivated and the point at which diminishing returns drive costs to an intolerable level. The aluminum assays for 3 different Lake Attitash samples provided similar results and indicated that diminishing benefits were encountered at doses higher than 30 g/m2 and that very little benefit was gained at doses >40 g/m2 (Figure 17). Ideally, we would like to drive Fe-P below 100 mg/kg and preferably below 50 mg/kg, but the reduction in internal load is roughly proportional to the percent reduction in Fe-P, so major reductions can be achieved even when Fe-P remains above detection.
While the laboratory aluminum assay is not a perfect replication of field conditions, these have proven accurate in other MA treatments. Also, treatments are believed to be additive; later application is not compromised by earlier application, so if more treatment is needed, it can be conducted later with no reduction in effectiveness. While it is preferable to treat once and be done with the process, overdosing raises costs without providing clear benefits. Extreme underdosing will not attain the desired benefits, but a dose slightly lower than optimal appears to only affect longevity of benefits, not the degree of initial improvement. Consequently, a lower dose with potential follow up treatment some years later becomes an attractive option when costs are high.
Diminishing benefits of treatment are a function of external loading, release of phosphorus from organic decay, and upward P migration through the treated sediment (Huser et al. 2016). External loading affects results for all lakes, and aluminum treatment is not advised when external loading is still the dominant P source. That does not appear to be the case for Lake Attitash based on significant storm water BMPs and EPA analysis, and aluminum treatment should provide relief from algae blooms for many years. Certainly there is ongoing external loading and further reductions will help protect the lake, but the efforts of the last two decades
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Attitash Sample #267 Aluminum Dosing and Inactivated Iron-Bound Phosphorus 900 800 700 600 500 400 300
Phosphorusmg/kg 200 100 0 0 20 40 60 80 100 Aluminum Dose g/m2
Attitash Sample #272 Aluminum Dosing and Inactivated Iron-Bound Phosphorus 1200
1000
800
600
400 Phosphorusmg/kg 200
0 0 20 40 60 80 100 Aluminum Dose g/m2
Attitash Composited Samples #269, 270 and 273 Aluminum Dosing and Inactivated Iron-Bound Phosphorus 400 350 300 250 200 150
Phosphorusmg/kg 100 50 0 0 20 40 60 80 100 Aluminum Dose g/m2
Figure 17. Aluminum Assay Results for Lake Attitash Sediment
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suggest that much progress has been made and that further reductions will be both difficult and expensive. Export coefficients for phosphorus for the 3 defined basins are near the low end for drainage areas with residential development and agricultural uses. Further, relatively long detention time suggests that any in-lake action will have more immediate benefits that will last for several years based on water column effects alone, and for one to two decades based on expected sediment effects.
For unstratified lakes that mix frequently, organic phosphorus is an important source, although phosphorus released from Fe-P during intermittent periods of anoxia can also be an important phosphorus source. Such lakes tend to exhibit improved conditions for about 10 years after proper aluminum treatment. Lake Attitash does not stratify strongly, but is more stratified with more prolonged anoxia at the sediment-water interface than lakes we would consider unstratified. Mixing in response to major storms or wind events is possible, but Lake Attitash tends to behave more like a standard stratified lake. That stratification breaks down more easily and earlier in late summer than for deeper lakes, but appears to set up by late June and be fairly stable in July and August. Decay of organic matter will provide some phosphorus in Lake Attitash, but Fe-P is likely to be the most critical source, especially during summer. Duration of benefits in excess of 10 years is therefore expected.
For lakes that more strongly stratify, the upward migration of Fe-P through the treatment zone is expected to be the dominant influence on duration of benefits. Duration of benefits has averaged about 20 years over a range of treatments in stratified lakes (Huser et al. 2016). Based on one core sample from Lake Attitash, the concentration of Fe-P below 10 cm is <220 mg/kg. While not low, this is about half the concentration in most of the targeted upper layer and suggests a slower rate of replenishment of surficial sediment Fe-P.
Speculated impact of powerboating an aluminum treatment effectiveness has largely been unwarranted. For Lake Attitash, with treatment in water >15 feet deep, any influence by prop wash is extremely unlikely. Boating impact on aluminum treatment effectiveness has not been documented even in shallower treatments where disturbance by motor use might be more rationally envisioned.
Considering the features of Lake Attitash and experience elsewhere, the benefits of an effective aluminum treatment should last between 15 and 20 years, possibly longer if external loading is further controlled and/or upward migration of Fe-P from deeper sediments is of the lower magnitude expected. If the dose turns out to be lower than optimal, the initial results should still be quite impressive, but they would not last as long. Current thinking in aluminum treatments is not to overdose, but to monitor sediment cores to assess the rate at which Fe-P is migrating upward to predict when additional treatment might be needed. Additional sampling is advised as part of any aluminum treatment project.
Applying the LLRM with a 90% reduction in internal loading yielded a predicted post-treatment average surface water phosphorus concentration of just under 15 ug/L and provided water clarity and a temporal chlorophyll distribution that should represent conditions acceptable to lake users and for secondary water supply. The aluminum assays depicted in Figure 16 suggest that the recommended dose of 40 g/m2 would result in a 86% reduction in iron-bound phosphorus
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availability for the first (top) sample, a 90% reduction for the second (middle) sample, and a 70% reduction for the third (bottom) sample. However, the sample that exhibited a 70% reduction represents a composite of most of the treatment area, so the overall internal load reduction may be closer to 70% than 90%. Re-running the LLRM with a 70% reduction in internal load results in a predicted post-treatment average phosphorus level of 16 ug/L, within the 15-17 ug/L range suggested as adequate to meet proposed target lake conditions.
The benefits of phosphorus inactivation may extend to oxygen improvements as well as a decrease in the available phosphorus. Most treatments have reduced but not eliminated anoxia in the deeper water. The reduction in algae translates to less oxygen-demanding organic matter settling into deeper water, but the ongoing oxygen demand of the organic sediments is not appreciably reduced by the treatment, so a thinner anoxic layer that forms later in the summer is often observed.
Low oxygen is known to occur at the sediment-water interface in winter over an area of about 194 acres, and may well occur over an area that large in summer; monitoring to detect anoxia at the sediment-water interface would involve lowering the probe right into the muck, something that is often not done in surveys. The 194 acre area corresponds to the 13 ft (4 m) depth contour, about the depth where regular mixing ceases in summer (Figure 10); this is a logical target treatment area.
Cost considerations cannot be ignored when planning aluminum treatments of this magnitude. A dose of 209 g/m2, as suggested by an Al:P ratio of 40:1 for the average Fe-P concentration over the maximum 194 acre target area, would cost close to $2 million. Alternatively, applying an Al:P ratio of 10:1 with the average Fe-P concentration leads to a dose of 52 g/m2; over a 194 acre treatment area, the estimated cost would be between $552,000 and $608,000. Using the 40 g/m2 dose suggested by aluminum assays, treatment of a 194 ac area would cost between $450,000 and $470,000. If funds are available to dose the 20 acres with higher Fe-P at 100 g/m2, that would add about $72,000 to the cost of the 40 g/m2 treatment and would be worthwhile if affordable.
Multiple combinations of dose and treated area could be considered to stay within the limits of available funding. No less than 130 acres, the documented anoxic area in July and August, should be treated. No less than 40 g/m2, the dose indicated by the aluminum assays, should be applied. This suggests a minimum treatment cost of about $320,000.
An alternative is to apply a lower dose of aluminum to the water column to strip phosphorus from it going into the summer, when little flushing occurs. The dose needs to be at least 20 times the phosphorus concentration, which averages about 44 ug/L in the deeper water in the summer, so a dose of at least 1 mg/L would be recommended. This equates to a cost of about $41,000 for a treatment. However, the stripping of phosphorus from the water column at lower concentrations than encountered in the surficial sediment is not very efficient; the resulting water column phosphorus concentration may not be as low as desired. So it is possible that a larger dose will be needed or that two treatments might be needed in a summer.
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The settled floc from water column treatment will inactivate some surficial sediment phosphorus, but the dose for the treatment described above is about 3.5 g/m2, so about 12 water column treatments would be needed to provide the recommended surficial sediment inactivation. The cost would approach $500,000, excluding inflation over the years it would take to add this dose.
Recommendations Based on the documented condition of Lake Attitash and current understanding of the factors contributing to that condition, the most important action that can be taken to improve the lake is the reduction of internal loading of phosphorus from sediments exposed to anoxia. This can be accomplished by multiple means, but considering cost, flexibility of implementation, and overall expected effectiveness, it is recommended that a single treatment of surficial sediments with aluminum over a 194 acre area below a water depth of 3.5 m (11.5 feet) be performed. The recommended dose is 40 g/m2, although a 20 acre area (the deepest part of the lake) could be treated at 100 g/m2. Such a treatment could be conducted at almost any time, but maximum effectiveness and most immediate improvement of the lake would be obtained with a spring treatment. This treatment should cost less than $600,000.
The proposed treatment is expected to reduce average phosphorus in the upper portion of the water column to about 16 ug/L, resulting in average water clarity of 3.0 m (10 feet) and algal chlorophyll levels that exceed 10 ug/L less than 10% of the time. No severe cyanobacteria blooms should occur. The duration of benefits is expected to be about 15 years, possibly up to 20 years. Oxygen in deep water should increase, although some anoxia is still expected from about mid-summer in the deepest area. Recreational lake use and suitability for water supply will be greatly enhanced, and the lake should then have water quality commensurate with its classification as a Class A Outstanding Resource Water. The increased clarity is likely to promote increase rooted plant growth, and plans should be made to apply physical or chemical means to address potential plant nuisances.
Additional watershed management efforts are warranted to limit external inputs, but extremely large expenditures to make slight reductions in external loading should not be pursued. If watershed loading is found to be sufficient to foster problems in the lake after treatment to reduce internal loading, consideration could be given to either inflow treatment (most likely for the Back River) or water column treatment with aluminum in mid- to late spring.
To document project success and provide data to aid additional management planning, a monitoring program such as had been conducted prior to 2011 or as recommended by the USEPA in its 2014 assessment report should be instituted. Aside from funding limitations, lack of data is the greatest impediment to ongoing management. An appropriate program could be conducted by volunteers with professional support for about $10,000 per year.
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References
AECOM. 2009. LLRM – Lake Loading Response Model: Users Guide and Quality Assurance Project Plan. AECOM, Willington, CT.
Amesbury, MA. 2006. Lake Attitash Stormwater Treatment Program. Sec 319 Project 01- 20/319. Prepared for the MA DEP and USEPA. Amesbury, MA.
Amesbury, MA. 2014. Lake Attitash Watershed Restoration. Sec 319 Project 11-07/319. Prepared for the MA DEP and USEPA. Amesbury, MA.
CDM. 1999. Lake Attitash Watershed Management Plan. CDM, Cambridge, MA.
Huser, B., S. Egemose, H. Harper, M. Hupfer, H. Jensen, K. Pilgrim, K. Reitzel, E. Rydin, M. Futter. 2016. Longevity and effectiveness of aluminum addition to reduce sediment phosphorus release and restore lake water quality. Water Res. (in press).
James, W. and J. Bischoff. 2015. Relationships between redox-sensitive phosphorus concentrations in sediment and the aluminum:phosphorus binding ratio. Lake Reserv. Manage. 31:339-346.
OWRB (Oklahoma Water Resources Board). 2012. Lake Thunderbird Water Quality 2011. Oklahoma City, Okla: OWRB.
Ruggirio, K. 2014. Where have all the Daphnia gone in Lake Attitash? UNH Center for Freshwater Biology Research 14(1): 1-4.
Snook, H. 2014. Lake Attitash Assessment Report. USEPA Region I Laboratory, Chelmsford, MA.
Wagner, K.J. 2001. In-Lake Management. Chapter 7 in Holdren, C., W. Jones, and J. Taggart. 2001. Managing Lakes and Reservoirs. North American Lake Management Society and Terrene Institute, EPA 841-B-01-006, Washington, DC.
Wagner, K. 2015. Oxygenation and Circulation as Aids to Water Supply Reservoir Management. Water Research Foundation, Denver, CO.
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