Following an algaecide treatment, applied copper moves from the water column to algae and ultimately to bottom sediments in an aquatic system [22] . The transport of copper to sediments is mediated by sorption to suspended particulates and precipitation via interaction with carbonate and sulfide [3] . Liu et al. [23] measured rapid and strongly bound copper sorption to suspended sediment particles within 2 hours after copper applications with 99% of applied copper transferring to bottom sediments in 2 days. Movement of copper as an algaecide to sediments is intensely facilitated by algae biomass. Copper algaecides rapidly sorb (~15 minutes) to numerous binding sites (carboxylic, sulfhydryl, phosphate groups, transport proteins, etc.) on algae [24] , and the movement of copper applied as an algaecide has been monitored with settling algae following treatment [25] . This is consistent with Jones et al. [26] finding the majority of copper binding in treated sediments short term after application was associated with the oxidizable fraction comprised primarily of algae biomass. Therefore, movement of copper out of a water resource treated with a copper pesticide is strongly dependent on the settling of algae and other suspended solids.

Copper from an algaecide or herbicide treatment is not expected to be measurable in bottom sediments where the application method only treats a portion of the water body (e.g. band treatments), due to dilution and lateral copper movement. A study by Iwinski et al. [27] comparing coves treated ~5 times a year for up to 10 years indicated that a majority of the sites had similar copper concentrations as adjacent untreated coves (p = 0.77, α = 0.05; treated n = 27, mean = 51.6 mg Cu/kg; untreated n = 27, mean = 151.3 mg Cu/kg). In circumstances of repeated or large treatment areas (e.g. whole water column treatment), copper from an application could potentially increase sediment copper concentration [28] . The increase of sediment copper concentration after an application, has been measured to range from 0.9 - 12 mg Cu/kg [28] [29] , respectively. The approximate tenfold increase between these two measured concentrations is hypothesized as a result of bioturbation and sediment accretion. For instance, the 0.9 mg Cu/kg was interpolated from the measured increase of sediment copper concentrations over three years with multiple applications in an aquaculture pond. While, the 12 mg Cu/kg was measured in the top 3 cm of sediment (6 mg Cu/kg in the top 8 cm) in a similar sized aquaculture pond soon after applied copper partitioned to the bottom sediments (2 - 3 weeks after treatment). This demonstrates copper is significantly diluted through time, which is likely attributed to accretion and copper migration to greater sediment depths. Liu et al. [23] measured copper to migrate up to 14 - 16 cm over a 12 week period thereby diluting the concentration. According to Jacinthe et al. [30] , the sediment copper concentration in three different central Indiana reservoirs increased by approximately 3 - 4 mg Cu/kg per algaecide application applied throughout four consecutive years of treatments. Hullebusch et al. [31] measured an increase of 7.7 mg Cu/kg following a copper sulfate application of 0.197 mg/L in a French reservoir with a 1.2 meter average depth. Overall, the expected increase of sediment copper concentration based on reported copper concentrations following applications is expected to be <8 mg Cu/kg per application. However, this accumulation is lessened through time due to downward migration and sediment accretion.

Copper that partitions to sediments following a pesticide application is transformed to more stable forms through time, such as malachite and chalcocite [23] . Jones et al. [26] found that the bioavailability of sediment-sorbed copper continues to transform to more stable and biologically unreactive forms following transport to sediments, which further decreases risks to biota. The sediment copper concentration that may cause adverse effects to non-target organisms can differ widely (orders of magnitude) depending on sediment characteristics [28] [32] . Acid volatile sulfides [33] , pH [34] , cation exchange capacity [35] , and organic matter content and type [36] [37] are important sediment characteristics that influence copper bioavailability in sediments. In context, a sandy and oxic sediment with low organic matter is expected to have the greatest fraction of bioavailable copper, while an anoxic high organic matter sediment primarily composed of fine particles (silt and clay) would have the least bioavailable fraction of copper. Overall, copper is a common constituent of the earth’s crust, with an ultimate fate in sediments and soils as relatively unreactive and thermodynamically stable forms.

Two common methods for screening sediments for copper toxicity are the acid volatile sulfides and simultaneously extracted metals ratio (AVS: SEM) and sediment quality guidelines [38] . The AVS:SEM technique is based on the amount of sulfides released from sediments simultaneously with metals using a 1 N HCl [39] , and a ratio greater than one indicates lack of toxicity [40] . However, this technique is limited to predicting the lack of toxicity, not presence, because AVS is not the only phase which copper partitions [38] . Values exist for screening sediments based on copper concentrations (NOAA Screening Quick Reference Tables) [41] , although these values predict potential for adverse effects and are likely conservative guidelines [22] , as well as lack site specificity [42] . Laboratory toxicity experiments have been used to measure the bioavailability of copper in sediments, and subsequently used to predict risks of copper in site specific sediments [22] [28] [32] . The bioavailability of copper is critical to assess as copper concentrations alone are often insufficient to accurately predict risks to in situ biota [35] . Laboratory experiments with site sediment amended with copper are considered to represent worst-case scenarios because there are no additional inputs of water or sediment and typically have a short acclimation period; these can reduce the bioavailability and risks of copper in sediments in the field [43] .

Overall, studies have demonstrated that sediments, from water bodies treated with copper, collected relatively shortly after treatment, as well as after years of copper pesticide treatments have not produced measurable adverse effects to sensitive sentinel species [22] [27] - [29] [32] . Specifically, Han et al. [29] found sediments from aquaculture ponds treated periodically for a series of years with copper sulfate to not elicit adverse effects to Hyalella azteca (amphipod) or Typha latifolia (common cattail). Huggett et al. [32] found the copper concentrations in sediments from a lake frequently treated with copper sulfate to control algae (Steilacoom Lake, WA, U.S.A.) were insufficient to elicit adverse effects to Hyalella azteca (amphipod), Chironomus tentans (infaunal midge), and Ceriodaphnia dubia (water flea). Gallagher et al. [22] found sediments from Lake Murray, SC, U.S.A., in areas annually treated with a chelated copper herbicide, did not produce adverse effects to Hyalella azteca or Ceriodaphnia dubia. Willis [28] found the copper bioavailability in sediments from an aquaculture pond collected three weeks following a chelated copper algaecide application was inadequate to elicit adverse effects to Hyalella azteca. Iwinski et al. [27] found that coves in a lake treated ~5 times a year for up to 10 years lacked measurable adverse effects based on laboratory toxicity experiments and in situ benthic abundance data.